World’s Soil Resources

Status of the 

World’s 

Main Report 

Soil Resources 

© FAO | Giuseppe Bizzarri 

INTERGOVERNMENTAL 

TECHNICAL PANEL ON SOILS

Status of the 

World’s Soil Resources 

Main report 

Prepared by 

Intergovernmental Technical Panel on Soils (ITPS) 

Luca Montanarella (Chair), Mohamed Badraoui, Victor Chude, Isaurinda Dos Santos Baptista Costa, 

Tekalign Mamo, Martin Yemefack, Milkha Singh Aulang, Kazuyuki Yagi, Suk Young Hong, Pisoot Vijarnsorn, 

Gan Lin Zhang, Dominique Arrouays, Helaina Black, Pavel Krasilnikov, Jaroslava Sobocá, Julio Alegre, Carlos 

Roberto Henriquez, Maria de Lourdes Mendonça-Santos, Miguel Taboada, David Espinosa Victoria, Abdullah 

Alshankiti, Sayed Kazem Alavi Panah, Elsiddig Ahmed El Mustafa El Sheikh, Jon Hempel, Dan Pennock, Marta 

Camps Arbestain, Neil McKenzie. 

Editorial board: Luca Montanarella (Chair), Victor Chude (Africa), Kazuyuki Yagi (Asia), Pavel Krasilnikov 

(Europe), Seyed Kazem Alavi Panah (Near East and North Africa), Maria de Lourdes Mendonça-Santos (Latin 

America and the Caribbean), Dan Pennock (North America), Neil McKenzie (SW Pacific). 

Managing editor: Freddy Nachtergaele 

Coordinating Lead Authors and Regional Coordinators/Authors: 

Mubarak Abdelrahman Abdalla, Seyed Kazem Alavipanah, André Bationo, Victor Chude, Juan Comerma, Maria 

Gerasimova, Jon Hempel, Srimathie Indraratne, Pavel Krasilnikov, Neil McKenzie, Maria de Lourdes Mendonça- 

Santos, Chencho Norbu, Ayo Ogunkunle, Dan Pennock, Thomas Reinsch, David Robinson, Pete Smith, Miguel 

Taboada and Kazuyuki Yagi. 

Reviewing Authors: Dominique Arrouays, Richard Bardgett, Marta Camps Arbestain, Tandra Fraser, Ciro 

Gardi, Neil McKenzie, Luca Montanarella, Dan Pennock and Diana Wall. 

Editorial team: Lucrezia Caon, Nicoletta Forlano, Cori Keene, Matteo Sala, Alexey Sorokin, Isabelle Verbeke, 

Christopher Ward. 

GSP Secretariat: Moujahed Achouri, Maryse Finka and Ronald Vargas. 

Other contributing Authors: 

Adams, Mary Beth 

Adhya Tapan, Kumar 

Agus, Fahmuddin 

Al Shankithi, Abdullah 

Alegre, Julio 

Aleman, Garcia 

Alfaro, Marta 

Alyabina, Irina 

Anderson, Chris 

Anjos, Lucia 

Arao, Tomohito 

Asakawa, Susumu 

Aulakh, Milkha 

Ayuke, Frederick 

Bai, Zhaohai 

Baldock, Jeff 

Balks, Megan 

Balyuk, Svyatoslav 

Bardgett, Richard 

Basiliko, Nathan 

Batkhishig, Ochirbat 

Bedard-Haughn, Angela 

Bielders, Charles 

Bock, Michael 

Bockheim, James 

Bondeau, Alberte 

Brinkman, Robert 

Bristow, Keith 

Broll, Gabrielle 

Bruulsma, Tom 

Bunning, Sally 

Bustamante, Mercedes 

Caon, Lucrezia 

Carating, Rodel 

Cerkowniak, Darrel 

Charzynski, Przemyslaw 

Clark, Joanna 

Clothier, Brent 

Coelho, Maurício Rizzato 

Colditz, Roland René 

Collins, Alison 

Compton, Jana 

Condron, Leo 

Corso, Maria Laura 

Cotrufo, Francesca 

Critchley, William 

Cruse, Richard 

da Silva, Manuela 

Dabney, Seth 

Daniels, Lee 

de Souza Dias, Moacir 

Status of the World’s Soil Resources | Main Report 

II

Dick, Warren 

Dos Santos Baptista, Isaurinda 

Drury, Craig 

El Mustafa El Sheikh, Ahmed 

Elsiddig 

Elder-Ratutokarua, Maria 

Elliott, Jane 

Espinosa, David 

Fendorf, Scott 

Ferreira, Gustavo 

Flanagan, Dennis 

Gafurova, Laziza 

Gaistardo, Carlos Cruz 

Govers, Gerard 

Grayson, Sue 

Griffiths, Robert 

Grundy, Mike 

Hakki Emrah, Erdogan 

Hamrouni, Heidi 

Hanly, James 

Harper, Richard 

Harrison, Rob 

Havlicek, Elena 

Hempel, Jon 

Henriquez, Carlos Roberto 

Hewitt, Allan 

Hiederer, Roland 

House, Jo 

Huising, Jeroen 

Ibánez, Juan José 

Jain, Atul 

Jefwa, Joyce 

Jung, Kangho 

Kadono, Atsunobu 

Kawahigashi, Masayuki 

Kelliher, Frank 

Kihara, Job 

Konyushkova, Maria 

Kuikman, Peter 

Kuziev, Ramazan 

Lai, Shawntine 

Lal, Rattan 

Lamers, John 

Lee, Dar-Yuan 

Lee, Seung Heon 

Lehmann, Johannes 

Leys, John 

Lobb, David 

Ma, Lin 

Macias, Felipe 

Maina, Fredah 

Mamo, Tekalign 

Mantel, Stephan 

McDowell, Richard 

Medvedev, Vitaliy 

Miyazaki, Tsuyushi 

Moore, John 

Morrison, John 

Mung'atu, Joseph 

Muniz, Olegario 

Nachtergaele, Freddy 

Nanzyo, Masami 

Ndiaye, Déthié 

Neall, Vince 

Noroozi, Ali Akbar 

Obst, Carl 

Ogle, Stephen 

Okoth, Peter 

Omutu, Christian 

Or, Dani 

Owens, Phil 

Pan, Genxing 

Panagos, Panos 

Parikh, Sanjai 

Pasos Mabel, Susana (†) 

Paterson, Garry 

Paustian, Keith 

Pietragalla, Vanina 

Pla Sentis, Ildefonso 

Polizzotto, Matthew 

Pugh, Thomas 

Qureshi, Asad 

Reddy, Obi 

Reid, D. Keith 

Richter, Dan 

Rivera-Ferre, Marta 

Rodriguez Lado, Luis 

Roskruge, Rick 

Rumpel, Cornelia 

Rys, Gerald 

Schipper, Louis 

Schoknecht, Noel 

Seneviratne, Sonia 

Shahid, Shabbir 

Sheffield, Justin 

Sheppard, Steve 

Sidhu, Gurjant 

Sigbert, Huber 

Smith, Scott 

Sobocká, Jaroslava 

Sönmez, Bülent 

Spicer, Anne 

Sposito, Garrison 

Stolt, Mark 

Suarez, Don 

Takata, Yusuke 

Tarnocai, Charles 

Tassinari, Diego 

Tien, Tran Minh 

Toth, Tibor 

Trumbore, Susan 

Tuller, Markus 

Urquiaga Caballero, Segundo 

Urquiza Rodrigues, Nery 

Van Liedekerke, Marc 

Van Oost, Kristof 

Vargas, Rodrigo 

Vargas, Ronald 

Vela, Sebastian 

Vitaliy, Medvedev 

Vrscaj, Boris 

Waswa, Boaz 

Watanabe, Kazuhiko 

Watmough, Shaun 

Webb, Mike 

Weerahewa, Jeevika 

West, Paul 

Wiese, Liesl 

Wilding, Larry 

Xu, Renkou 

Yan, Xiaoyuan 

Yemefack, Martin 

Yokoyama, Kazunari 

Zhang, Fusuo 

Zhou, Dongme i 

Zobeck, Ted 

FOOD AND AGRICULTURE ORGANIZATION OF THE UNITED NATIONS 

Rome, 2015 


III

Disclaimer and copyright 

Recommended citation: 

FAO and ITPS. 2015. 

Status of the World’s Soil Resources (SWSR) – Main Report. 

Food and Agriculture Organization of the United Nations 

and Intergovernmental Technical Panel on Soils, Rome, Italy 

The designations employed and the presentation of material in this information product do not imply the 

expression of any opinion whatsoever on the part of the Food and Agriculture Organization of the United 

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ISBN 978-92-5-109004-6 

© FAO, 2015 

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IV

Table of contents 

Disclaimer and copyright | IV 

Table of contents | V 

Foreword | XIX 

Preface | Scope of The State of the World’s Soil Resources | XXI 

Acknowledgment | XXII 

List of abbreviations | XXIV 

List of tables | XXXIII 

List of boxes | XXXIV 

List of figures | XXXV 

Preface | 1 

Global soil resources | 3 

1 | Introduction | 4 

1.1 | The World Soil Charter | 4 

1.2 | Basic concepts | 7 

Sustainable soil management | 8 

Soil degradation and threats to soil functions | 8 

Soil functions and ecosystem services | 9 

Soils and natural capital | 9 

Planetary boundaries and safe operating space for humanity | 9 

Biodiversity | 10 

2 | The role of soils in ecosystem processes | 13 

2.1 | Soils and the carbon cycle | 13 

2.1.1 | Quantitative amounts of organic C stored in soil | 14 


V

2.1.2 | Nature and formation of soil organic C | 15 

2.1.3 | Soil C pools | 16 

2.1.4 | Factors influencing soil C storage | 17 

2.1.5 | Carbon cycle: knowledge gaps and research needs | 18 

2.1.6 | Concluding remarks | 18 

2.2 | Soils and the nutrient cycle | 18 

2.2.1 | The nutrient cycle: knowledge gaps and research needs | 21 

2.3 | Soils and the water cycle | 21 

2.4 | Soil as a habitat for organisms and a genetic pool | 24 

3 | Global Soil Resources | 31 

3.1 | The evolution of soil definitions | 31 

3.2 | Soil definitions in different soil classification systems | 32 

3.3 | Soils, landscapes and pedodiversity | 32 

3.4 | Properties of the soil | 33 

3.5 | Global soil maps | 33 

3.6 | Soil qualities essential for the provision of ecosystem services | 34 

3.6.1 | Inherent soil fertility | 35 

3.6.2 | Soil moisture qualities and limitations | 37 

3.6.3 | Soils properties and climate change | 37 

3.6.4 | Soil erodibility and water erosion | 38 

3.6.5 | Soil workability | 39 

3.6.6 | Soils and ecosystem goods and services | 40 

3.7 | Global assessments of soil change - a history | 43 

3.7.1 | GLASOD: expert opinion | 43 

3.7.2 | LADA-GLADIS: the ecosystem approach | 45 

3.7.3 | Status of the World’s Soil Resources | 46 

4 | Soils and Humans | 50 


VI

4.1 | Current land cover and land use | 50 

4.2 | Historical land cover and land use change | 53 

4.3 | Interactions between soils, land use and management | 54 

4.3.1 | Land use change and soil degradation | 54 

4.3.2 | Land use intensity change | 60 

4.3.3 | Land use change resulting in irreversible soil change | 65 

4.4 | Atmospheric deposition | 72 

4.4.1 | Atmospheric deposition | 72 

4.4.2 | Main atmospheric pollutants: Synopsis of current state of knowledge | 73 

4.4.3 | Knowledge gaps and research needs | 76 

Global Soil Change Drivers, Status and Trends | 88 

5 | Drivers of global soil change | 89 

5.1 | Population growth and urbanization | 89 

5.1.1 | Population dynamics | 89 

5.1.2 | Urbanization | 91 

5.2 | Education, cultural values and social equity | 91 

5.3 | Marketing land | 92 

5.4 | Economic growth | 94 

5.5 | War and civil strife | 94 

5.6 | Climate change | 96 

6 | Global soil status, processes and trends | 100 

6.1.1 | Processes | 100 

6.1.2 | Status of Soil Erosion | 101 

6.1.3 | Soil erosion versus soil formation | 103 

6.1.4 | Soil erodibility | 104 

6.1.5 | Soil erosion and agriculture | 104 

6.1.6 | Soil erosion and the environment | 105 


VII

6.1.7 | Effects of hydrology and water | 106 

6.1.8 | Vegetation effects | 107 

6.1.9 | Alteration of nutrient and dust cycling | 107 

6.1.10 | Trends in soil erosion | 108 

6.1.11 | Conclusions | 108 

6.2 | Global soil organic carbon status and trends | 109 

6.2.1 | Introduction | 109 

6.2.2 | Estimates of global soil organic carbon stocks | 109 

6.2.3 | Spatial distribution of SOC | 111 

6.2.4 | Spatial distribution of carbon in biomass | 113 

6.2.5 | Distribution of terrestrial carbon pool by vegetation class | 114 

6.2.6 | Historic trends in soil carbon stocks | 116 

6.2.7 | Future loss of SOC under climate change | 118 


6.3 | Soil contamination status and trends | 119 


6.3.2 | Global status of soil contamination | 119 

6.3.3 | Trends and legislation | 121 

6.4 | Soil acidification status and trends | 122 

6.4.1 | Processes and causes of acidification | 122 

6.4.2 | Impact of soil acidification | 123 

6.4.3 | Responses to soil acidification | 123 

6.4.4 | Global status and trends of soil acidification | 123 

6.5 | Global status of soil salinization and sodification | 124 

6.5.1 | Status and extent | 124 

6.5.2 | Causes of soil salinity | 126 

6.5.4 | Trends and impacts | 126 


VIII

6.5.5 | Responses | 126 

6.6 | Soil biodiversity status and trends | 127 


6.6.2 | Soil biota and land use | 128 


6.7 | Soil sealing: status and trends | 130 

6.8 | Soil nutrient balance changes: status and trends | 132 


6.8.2 | Principles and components of soil nutrient balance calculations | 133 

6.8.3 | Nutrient budgets: a matter of spatial scale | 134 

6.8.4 | Nutrient budgets: a matter of land use system, land use type, 

managementand household equity | 135 

6.8.5 | What does the future hold? | 136 

6.9 | Soil compaction status and trends | 137 

6.9.1 | Effect of tillage systems on compaction | 138 

6.9.2 | What is the extent of deep soil compaction? | 139 

6.9.3 | Solutions to soil compaction problems | 139 

6.10 | Global soil-water quantity and quality: status, processes and trends | 140 

6.10.1 | Processes | 140 

6.10.2 | Quantifying soil moisture | 142 

6.10.3 | Status and trends | 143 

6.10.4 | Hotspots of pressures on soil moisture | 144 


Soil change: impacts and responses | 168 

7 | The impact of soil change on ecosystem services | 169 

7.1 | Introduction | 169 

7.2 | Soil change and food security | 172 


IX

7.2.1 | Soil erosion | 175 

7.2.2 | Soil sealing | 178 

7.2.3 | Soil contamination | 178 

7.2.4 | Acidification | 178 

7.2.5 | Salinization | 178 

7.2.6 | Compaction | 179 

7.2.7 | Nutrient imbalance | 179 

7.2.8 | Changes to soil organic carbon and soil biodiversity | 179 

7.3 | Soil change and climate regulation | 181 

7.3.1 | Soil carbon | 181 

7.3.2 | Nitrous oxide emissions | 183 

7.3.3 | Methane emissions | 184 

7.3.4 | Heat and moisture transfer | 185 

7.4 | Air quality regulation | 188 

7.4.2 | Ammonia emissions | 188 

7.4.3 | Aerosols | 188 

7.5 | Soil change and water quality regulation | 189 

7.5.1 | Nitrogen and phosphorous retention and transformation | 190 

7.5.2 | Acidification buffering | 191 

7.5.3 | Filtering of reused grey water | 192 

7.5.4 | Processes impacting service provision | 192 

7.6 | Soil change and water quantity regulation | 194 

7.6.2 | Precipitation interception by soils | 194 

7.6.3 | Surface water regulation | 195 

7.7 | Soil change and natural hazard regulation | 195 

7.7.1 | Soil landslide hazard | 197 

7.7.2 | Soil hazard due to earthquakes | 198 


X

7.7.3 | Soil and drought hazard | 198 

7.7.4 | Soil and flood hazard | 199 

7.7.5 | Hazards induced by thawing of permafrost soil | 199 

7.8 | Soil biota regulation | 199 

7.9 | Soils and human health regulation | 201 

7.10 | Soil and cultural services | 203 

8 | Governance and policy responses to soil change | 223 

8.2 | Soils as part of global natural resources management | 224 

8.2.1 | Historical context | 224 

8.2.2 | Global agreements relating to soils | 225 

8.3 | National and regional soil policies | 228 

8.3.1 | Sustainable soil management – criteria and supporting practices | 228 

8.3.2 | Education about soil and land use | 229 

8.3.3 | Soil research, development and extension | 229 

8.3.4 | Private benefits, public goods and payments for ecosystem services | 229 

8.3.5 | Intergenerational equity | 230 

8.3.6 | Land degradation and conflict | 230 

8.4 | Regional soil policies | 231 

8.4.1 | Africa | 231 

8.4.2 | Asia | 232 

8.4.3 | Europe | 232 

8.4.4 | Eurasia | 232 

8.4.5 | Latin America and the Caribbean (LAC) | 233 

8.4.6 | The Near East and North Africa (NENA) | 233 

8.4.7 | North America | 234 

8.4.8 | Southwest Pacific | 234 


XI

8.5 | Information systems, accounting and forecasting | 235 

8.5.1 | Soil information for markets | 236 

8.5.2 | Environmental accounting | 236 

8.5.3 | Assessments of the soil resource | 237 

9 | Regional Assessment of Soil Changes in Africa South of the Sahara | 242 


9.2 | Stratification of the Region | 244 

9.2.1 | Arid zone | 244 

9.2.2 | Semi-arid zone | 246 

9.2.3 | Sub-humid zone | 246 

9.2.5 | Highlands zone | 247 

9.3 | General soil threats in the region | 247 

9.3.1 | Erosion by water and wind | 247 

9.3.2 | Loss of soil organic matter | 248 

9.3.3 | Soil nutrient depletion | 249 

9.3.4 | Loss of soil biodiversity | 250 

9.3.5 | Soil contamination and pollution | 251 

9.3.6 | Soil acidification | 252 

9.3.7 | Salinization and sodification | 252 

9.3.8 | Waterlogging | 252 

9.3.9 | Compaction, crusting and sealing | 252 

9.4 | The most important soil threats in Sub-Saharan Africa | 253 


9.4.2 | Loss of soil organic matter | 258 

9.4.3 | Soil nutrient depletion | 260 

9.5 | Case studies | 263 

9.5.1 | Senegal | 263 


XII

9.5.2 | South Africa | 266 

9.6 | Summary of conclusions and recommendations | 275 

10 | Regional Assessment of Soil Change in Asia | 287 


10.2. Stratification of the region | 288 

10.2. 1 | Climate and agro-ecology | 288 

10.2.2 | Previous regional soil assessments | 289 

10.3 | General threats to soils in the region | 291 

10.3.1 | Erosion by wind and water | 291 

10.3.2 | Soil organic carbon change | 291 



10.3.5 | Soil salinization and sodification | 293 




10.3.9 | Compaction | 296 

10.3.10 | Sealing and capping | 297 

10.4 | Major threats to soils in the region | 297 

10.4.1 | Erosion | 297 


10.4.3 | Soil salinization and sodification | 301 

10.4.4 | Nitrogen imbalance | 302 


10.5.1 | Case study for India | 304 

10.5.2 | Case study for Indonesia | 307 

10.5.3 | Case study for Japan | 310 


XIII

10.5.4 | Case study of greenhouse gas emissions from paddy fields | 314 

10.6 | Conclusion | 315 

11 | Regional assessment of soil changes in Europe and Eurasia | 330 


11.2 | Stratification of the region | 331 

11.3 | General threats to soils in the region | 335 

11.4. Major threats to soils in Europe and Eurasia | 338 



11.4.3 | Soil organic matter decline | 339 



11.5.1 | Case study: Austria | 344 

11.5.2 | Case study: Ukraine | 350 

11.5.3 | Case study: Uzbekistan | 353 

11.6 | Conclusion | 356 

12 | Regional assessment of soil changes in Latin America and the Caribbean | 364 


12.2 | Biomes, ecoregions and general soil threats in the region. | 366 

12.3. General soil threats in the region | 371 






12.3.6 | Compaction | 373 



XIV



12.3.10 | Sealing | 373 

12.4 | Major threats to soils | 374 



12.4.3 | Soil salinization | 380 


12.5.1 | Argentina | 382 

12.5.2 | Cuba | 386 

12.6 | Conclusions and recommendations | 388 

13 | Regional Assessment of Soil Changes in the Near East and North Africa | 399 


13.2 | Major land use systems in the Near East and North Africa | 402 

13.3 | Major threats to soils in the region | 404 

13.3.1 | Erosion | 404 




13.3.5 | Soil salinization/sodification | 407 


13.3.7 Waterlogging | 408 

13.3.8 | Nutrient balance change | 408 

13.3.9 | Compaction | 409 

13.3.10 | Sealing/capping | 409 

13.4 | Major soil threats in the region | 411 

13.4.1 | Water and wind erosion | 411 


XV

13.4.2 | Soil salinization/sodification | 416 




13.5.1 | Case study: Iran | 423 

13.6.2 | Case Study: Tunisia | 426 

13.6 | Conclusions | 430 

14 | Regional Assessment of Soil Changes in North America | 442 


14.2 | Regional stratification and soil threats | 443 

14.2.1 | Regional stratification and land cover | 443 

14.3 | Soil threats | 447 



14.3.3 | Soil salinization | 450 

14.3.4 | Soil sealing/capping | 452 

14.3.5 | Soil compaction | 453 

14.3.6 | Waterlogging and wetlands | 454 

14.4 | Major soil threats | 454 




14.4.4 | Soil biodiversity | 459 

14.5 | Case study: Canada | 460 

14.5.1 | Water and wind erosion | 460 



XVI


14.6 | Conclusions and recommendations | 467 

15 | Regional Assessment of Soil Change in the Southwest Pacific | 476 


15.2 | The major land types in the region | 477 

15.3 | Climate | 480 

15.4 | Land use | 480 

15.4.1 | Historical context | 480 

15.4.2 | Nineteenth and twentieth centuries | 481 

15.4.3 | Contemporary land-use dynamics | 482 

15.5 | Threats to soils in the region | 485 

15.5.1 | Erosion by wind and water | 485 








15.5.7 | Compaction | 497 



15.6.1 | Case study one: Intensification of land use in New Zealand | 498 

15.6.3 | Case study two: Soil management challenges in southwest Western 

Australia | 500 

15.6.2 | Case study three: Atoll Islands in the Pacific | 504 

15.6.4 | Case study four: DustWatch – an integrated response to wind erosion in 


XVII

Australia | 505 

15.7 | Conclusions | 507 

16 | Regional Assessment of Soil Change in Antarctica | 520 

16.1 | Antarctic soils and environment | 521 

16.2 | Pressures/threats for the Antarctic soil environment | 521 

16.3 | Response | 523 

Annex | Soil groups, characteristics, distribution and ecosystem services | 527 

1 | Soils with organic layers | 528 

2 | Soils showing a strong human influence | 530 

3 | Soils with limitations to root growth | 534 

4 | Soils distinguished by Fe/Al chemistry | 544 

5 | Soils with accumulation of organic matter in the topsoil | 561 

6 | Soils with accumulation of moderately soluble salts | 569 

7 | Soils with a clay-enriched subsoil | 575 

8 | Soils with little or no profile development | 585 

9 | Permanently flooded soils | 593 

Glossary of technical terms | 599 

Authors and affiliations | 602 


XVIII

Foreword 

This document presents the first major global assessment ever on soils and related issues. 

Why was such an assessment not carried out before? We have taken soils for granted for a long time. 

Nevertheless, soils are the foundation of food production and food security, supplying plants with nutrients, 

water, and support for their roots. Soils function as Earth’s largest water filter and storage tank; they contain 

more carbon than all above-ground vegetation, hence regulating emissions of carbon dioxide and other 

greenhouse gases; and they host a tremendous diversity of organisms of key importance to ecosystem 

processes. 

However, we have been witnessing a reversal in attitudes, especially in light of serious concerns expressed by 

soil practitioners in all regions about the severe threats to this natural resource. In this more auspicious context, 

when the international community is fully recognizing the need for concerted action , the Intergovernmental 

Technical Panel on Soils (ITPS), the main scientific advisory body to the Global Soil Partnership (GSP) hosted by 

the Food and Agriculture Organization of the United Nations (FAO), took the initiative to prepare this much 

needed assessment. 

The issuance of this first “Status of the World’s Soil Resources” report was most appropriately timed with the 

occasion of the International Year of Soils (2015) declared by the General Assembly of the United Nations. It was 

made possible by the commitment and contributions of hosts of reputed soil scientists and their institutions. 

Our gratitude goes to the Lead Authors, Contributing Authors, Editors and Reviewers who have participated 

in this effort, and in particular to the Chairperson of the ITPS, for his dedicated guidance and close follow up. 

Many governments have supported the participation of their resident scientists in the process and 

contributed resources, thus also assuring the participation of experts from developing countries and countries 

with economies in transition. In addition, a Technical Summary was acknowledged by representatives of 

governments assembled in the Plenary Assembly of the GSP, signaling their appreciation of the many potential 

uses of the underlying report. Even more comprehensive and inclusive arrangements will be sought in the 

preparations of further, updated versions. 

The report is aimed at scientists, laymen and policy makers alike. It provides in particular an essential 

benchmark against periodical assessment and reporting of soil functions and overall soil health at global 

and regional levels. This is of particular relevance to the Sustainable Development Goals (SDGs) that the 

international community pledged to achieve. Indeed, these goals can only be achieved if the crucial natural 

resources – of which soils is one – are sustainably managed. 

The main message of this first edition is that, while there is cause for optimism in some regions, the majority 

of the world’s soil resources are in only fair, poor or very poor condition. Today, 33 percent of land is moderately 

to highly degraded due to the erosion, salinization, compaction, acidification and chemical pollution of soils. 

Further loss of productive soils would severely damage food production and food security, amplify food-price 

volatility, and potentially plunge millions of people into hunger and poverty. But the report also offers evidence 

that this loss of soil resources and functions can be avoided. Sustainable soil management, using scientific 

and local knowledge and evidence-based, proven approaches and technologies, can increase nutritious food 

supply, provide a valuable lever for climate regulation and safeguarding ecosystem services. 


XIX

We can expect that the extensive analytical contents of this report will greatly assist in galvanizing action 

at all levels towards sustainable soil management, also in line with the recommendations contained in the 

updated World Soil Charter and as a firm contribution to achieve the Sustainable Development Goals. 

We are proud to make this very first edition of the Status of the World’s Soil Resources report available for 

the international community, and reiterate once again our commitment to a world free of poverty, hunger 

and malnutrition. 

JOSÉ GRAZIANO DA SILVA 

FAO Director-General 


XX

Preface | Scope of The State of the 

World’s Soil Resources 

The main objectives of The State of the World’s Soil Resources are: (a) to provide a global scientific 

assessment of current and projected soil conditions built on regional data analysis and expertise; (b) to 

explore the implications of these soil conditions for food security, climate change, water quality and quantity, 

biodiversity, and human health and wellbeing; and (c) to conclude with a series of recommendations for action 

by policymakers and other stakeholders. 

The book is divided into two parts. The first part deals with global soil issues (Chapters 1 to 8). This is 

followed by a more specific assessment of regional soil change, covering in turn Africa South of the Sahara, 

Asia, Europe, Latin America and the Caribbean, the Near East and North Africa, North America, the Southwest 

Pacific and Antarctica. (Chapters 9 to 16). The technical and executive summaries are published separately. 

In Chapter 1 the principles of the World Soil Charter are discussed, including guidelines for stakeholders to 

ensure that soils are managed sustainably and that degraded soils are rehabilitated or restored. For long, soil 

was considered almost exclusively in the context of food production. However, with the increasing impact of 

humans on the environment, the connections between soil and broader environmental concerns have been 

made and new and innovative ways of relating soils to people have begun to emerge in the past two decades. 

Societal issues such as food security, sustainability, climate change, carbon sequestration, greenhouse gas 

emissions, and degradation through erosion and loss of organic matter and nutrients are all closely related 

to the soil resource. These ecosystem services provided by the soil and the soil functions that support these 

services are central to the discussion in the report. 

In Chapter 2 synergies and trade-offs are reviewed, together with the role of soils in supporting ecosystem 

services, and their role in underpinning natural capital. The discussion then covers knowledge - and knowledge 

gaps - on the role of soils in the carbon, nitrogen and water cycles, and on the role of soils as a habitat for 

organisms and as a genetic pool. This is followed in Chapter 3 by an overview of the diversity of global soil 

resources and of the way they have been assessed in the past. Chapter 4 reviews the various anthropogenic 

and natural pressures - in particular, land use and soil management – which cause chemical, physical and 

biological variations in soils and the consequent changes in environmental services assured by those soils. 

Land use and soil management are in turn largely determined by socio-economic conditions. These 

conditions are the subject of Chapter 5, which discusses in particular the role of population dynamics, market 

access, education and cultural values as well as the wealth or poverty of the land users. Climate change and 

its anticipated effects on soils are also discussed in this chapter. 

Chapter 6 discusses the current global status and trends of the major soil processes threatening ecosystem 

services. These include soil erosion, soil organic carbon loss, soil contamination, soil acidification, soil 

salinization, soil biodiversity loss, soil surface effects, soil nutrient status, soil compaction and soil moisture 

conditions. 

Chapter 7 undertakes an assessment of the ways in which soil change is likely to impact on soil functions 

and the likely consequences for ecosystem service delivery. Each subsection in this chapter outlines key soil 

processes involved with the delivery of goods and services and how these are changing. The subsections 

then review how these changes affect soil function and the soil’s contribution to ecosystem service delivery. 

The discussion is organized according to the reporting categories of the Millennium Ecosystem Assessment, 

including provisioning, supporting, regulating and cultural services. 


XXI

Chapter 8 of the report explores policy, institutional and land use management options and responses to 

soil changes that are available to governments and land users. 

The regional assessments in Chapters 9 to 16 follow a standard outline: after a brief description of the main 

biophysical features of each region, the status and trends of each major soil threat are discussed. Each chapter 

ends with one or more national case studies of soil change and a table summarizing the results, including the 

status and trends of soil changes in the region and related uncertainties. 


XXII

Acknowledgments 

The Status of the World’s Soil Resources report was made possible by the commitment and voluntary work 

of the world’s leading soil scientists and the institutions they are affiliated with. We would like to express 

our gratitude to all the Coordinating Lead Authors, Lead Authors, Contributing Authors, Review Editors 

and Reviewers. We would also like to thank the editorial staff and the GSP Secretariat for their dedication in 

coordinating the production of this first seminal report. 

Appreciation is expressed to many Governments who have supported the participation of their resident 

scientists in this major enterprise. In particular, our gratitude to the European Commission who financially 

supported the development and publication of this report. 


XXIII

List of abbreviations 

AAFC 

ACSAD 

AD 

AEZ 

AFES 

AFSIS 

AGES 

AGRA 

AKST 

ALOS 

AMA 

AMF 

ANC 

AOAD 

AOT 

APO-FFTC 

ARC 

ASGM 

ASI 

ASP 

ASSOD 

AU 

BASE 

BC 

BD 

BDP 

BIH 

BMLFUW 

BNF 

BOM 

Agriculture and Agri-Food Canada 

Arab Centre for the Study of Arid Zones and Dry Lands 

Anno Domini 

Agro-Ecological Zones 

Association Française Pour L’étude Du Sol 

African Soil Information Service 

Austrian Agency for Health and Food Safety 

Alliance for a Green Revolution in Africa 

Agricultural Knowledge Science and Technology 

Advanced Land Observation Satellite 

Agencia De Medio Ambiente 

Arbuscular Mycorrhizal Fungi 

Acid-Neutralising Capacity 

Arab Organization for Agricultural Development 

Aerosol Optical Thickness 

Asian Productivity Organization- Food & Fertilizer Technology Center 

Agricultural Research Council 

Artisanal and Small-Scale Gold Mining 

Advanced Science Institutesseries 

Asia Soil Partnership 

Assessment of Human-Induced Soil Degradation in South and Southeast Asia 

African Union 

Biome of Australia Soil Environments 

(1) Black Carbon; (2) Before Christ 

Biodiversity 

Bureau for Development Policy 

Bosnia And Herzegovina 

Austrian Federal Ministry of Agriculture, Forestry, Environment and Water Management 

Biological Nitrogen Fixing 

Bureau of Meteorology 

BP Before Present (1 January 1950) 

C:N 

CA 

CAAA 

CAADP 

Carbon To Nitrogen Ratio 

Conservation Agriculture 

Clean Air Act Amendments 

Comprehensive Africa Agriculture Development Programme 


XXIV

CACILM 

CAMRE 

CAZRI 

CBD 

CBM-CFS 

CCAFS 

CCME 

CE 

CEC 

CECS 

CEPAL 

CF 

CGIAR 

CIAT 

CIFOR 

CITMA 

CLIMSOIL 

CLM 

Central Asian Countries Initiative for Land Management 

Council of Arab Ministers Responsible For the Environment 

Central Arid Zone Research Institute 

Convention on Biological Diversity 

Carbon Budget Model of the Canadian Forest Sector 

Climate Change, Agriculture and Food Security 

Canadian Council Of Ministers of the Environment 

Common Era (Also Current era or Christian era) 

(1) Cation Exchange Capacity; (2) Commission of the European Communities 

Chemicals of Emerging Concern 

Comisión Económica Para América Latina Y El Caribe 

Commercial Farming 

Global Agricultural Research Partnership 

International Center for Tropical Agriculture 

Center for International Forestry Research 

Ministerio De Ciencia, Tecnologia Y Medio Ambiente 

Review of Existing Information on the Interrelations between Soil and Climate Change 

Contaminated Land Management 

CMIP 5 Coupled Model Intercomparison Project Phase 5 

COM 

CONABIO 

CONAFOR 

COSMOS 

CRC 

CRP 

CSA 

CSIF-SLM 

CSIRO 

CSM-BGBD 

CSSRI 

CSWCR&TI 

DAFWA 

DBC 

DDT 

DEA 

DECA 

DED 

DENR 

Commission Working Documents 

Comision Nacional Para El Conocimiento Y Uso De La Biodiversidad 

Comisión Nacional Forestal 

Cosmic-Ray Soil Moisture Observing System 

Risk of Colorectal Cancer 

Conservation Reserve Program 

Climate-Smart Agriculture 

Country Strategic Investment Framework for Sustainable Land Management 

Commonwealth Scientific and Industrial Research Organisation 

Conservation and Sustainable Management of Below-Ground Biodiversity 

The Central Soil Salinity Research Institute 

Central Soil & Water Conservation Research & Training Institute (India) 

Department Of Agriculture and Food, Western Australia 

Dissolved Black Carbon 

Dichlorodiphenyltrichloroethane 

Deliberate Evacuation Area 

Department Of Environment and Conservation, Australia 

Dust Event Days 

Department Of Environment and Natural Resources 


XXV

DEST 

DGVMS 

DIC 

DLDD 

DNA 

DOC 

DOI 

DPYC 

DSEWPAC 

DSI 

DSMW 

EA-20km 

EAD 

EC DG ENV 

EC 

EEA 

EEAA 

EEZ 

ELD 

EM-DAT 

ENSO 

EOLSS 

EPA CERCLIS 

EPA 

ERW 

ES 

ESA 

ESAFS 

ESCWA 

ESDB 

ESP 

ESRI 

EU SCAR 

EU 

FAO 

FAOSTAT 

FAO-WRB 

Australian Government Department of Education, Science and Training 

Dynamic Global Vegetation Models 

Dissolved Inorganic Carbon 

Desertification, Land Degradation and Drought 

Deoxyribonucleic Acid 

Dissolved Organic Carbon 

Digital Object Identifier 

Dissolved Pyrogenic Carbon 

Department Of Sustainability, Environment, Water, Population and Communities 

Dust Storm Index 

Digital Soil Map of the World 

Twenty Km Evacuation Area 

Environment Agency Abu Dhabi 

European Commission Directorate-General for Environment 

European Commission 

European Environment Agency 

Egyptian Environmental Affairs Agency 

Exclusive Economic Zone 

Economics of Land Degradation 

Emergency Events Database 

El Niño Southern Oscillation 

Encyclopedia of Life Support Systems 

United States Environmental Protection Agency, Comprehensive Environmental 

Response, Contamination and Liability Information System 

United States Environmental Protection Agency 

Explosive Remnants of War 

Ecosystem Services 

United Nations Economic and Social Affairs Department 

East and Southeast Asia Federation of Soil Science Societies 

United Nations Economic and Social Commission for Western Asia 

European Soil Database 

Exchangeable Sodium Percentage 

Environmental Systems Research Institute 

European Standing Committee on Agricultural Research 

European Union 

Food and Agriculture Organization of the United Nations 

Food and Agriculture Organization Corporate Statistical Database 

Food and Agriculture Organization World Reference Base 


XXVI

FDNPS 

FFS 

FIA 

FSI 

FSR 

GAP 

GDP 

GEF 

GEO 

GHG 

GIS 

GIZ 

GLADA 

GLADIS 

GLASOD 

GLC 2000 

GLC-SHARE 

GLRD 

GRACE 

GRID 

GSBI 

GSM 

GSP 

HORTNZ 

HTAP 

HWSD 

HYDE 

IAASTD 

IAATO 

ICAR 

ICARDA 

ICBA 

ICBL 

ICRAF 

IDP 

IFA 

IFAD 

Fukushima Dai-Ichi Nuclear Power Station 

Farmer Field School 

Forest Inventory and Analysis 

Forest Survey in India 

Fund-Service-Resources 

Southeast Anatolia Development Project Region 

Gross Domestic Product 

Global Environment Facility 

Global Environmental Outlook 

Greenhouse Gases 

Geographic Information System 

Deutsche Gesellschaft Für Internationale Zusammenarbeit (GIZ) Gmbh 

Global Land Degradation Assessment 

Global Land Degradation Information System 

Global Assessment of Human-Induced Soil Degradation 

Global Land Cover 2000 Project 

Global Land Cover SHARE 

Gender and Land Rights Database 

Gravity Recover and Climate Experiment 

Global Resource Information Database 

Global Soil Biodiversity Initiative 

Global Soil Map 

Global Soil Partnership 

Horticulture New Zealand 

Hemispheric Transport of Air Pollution 

Harmonized World Soil Database 

History Database of the Global Environment 

International Assessment of Agricultural Knowledge, Science and Technology for 

Development 

International Association of Antarctic Tour Operators 

Indian Council of Agricultural Research 

International Center for Agriculture Research In The Dry Areas 

International Center for Biosaline Agriculture 

International Campaign to Ban Landmines 

International Center for Research in Agroforestry 

Internally Displaced Peoples 

International Fertilizers Association 

International Fund for Agricultural Development 


XXVII

IFADATA 

IFPRI 

IGT-AMA 

IIASA 

ILCA 

IMAGE 

IMBE 

IMF 

IMK-IFU 

INIA 

IPBES 

IPCC 

IRENA 

IROWC-N 

IROWC-P 

ISA 

ISAM 

ISBN 

ISCO 

ISCW 

ISFM 

ISO 

ISRIC 

ISS-CAS 

ISSS 

ITPS 

IUSS 

IW 

JRC 

LAC 

LADA 

LCCS 

LD 

LDCS 

LPFN 

LPJ-GUESS 

LRTAP 

LS 

International Fertilizer Industry Association Database 

International Food Policy Research Institute 

Instituto De Geografía Tropical Y La Agencia De Medio Ambiente 

International Institute for Applied Systems Analysis 

International Livestock Centre for Africa 

Integrated Modelling Of Global Environmental Change 

Mediterranean Istitute of Biodiversity and Ecology 

International Monetary Fund 

Institute of Meteorology and Climate Research Atmospheric Environmental Research 

Instituto De Investigaciones Agropecuarias (Chile) 

Intergovernmental Panel on Biodiversity and Ecosystem Services 

Intergovernmental Panel on Climate Change 

International Renewable Energy Agency 

The Indicator of Risk of Water Contamination by Nitrogen 

Indicator of Risk of Water Contamination by Phosphorus 

Impervious Surface Area 

Integrated Impacts of Climate Change Model 

International Standard Book Number 

International Soil Conservation Organization 

Institute for Soil, Climate and Water 

Integrated Soil Fertility Management 

International Standards Organization 

International Soil Reference and Information Centre 

Institute of Soil Science – Chinese Academy of Sciences 

International Society for the Systems Sciences 

Intergovernmental Technical Panel on Soils 

International Union of Soil Sciences 

International Waters 

Joint Research Centre (European Commission) 

Latin America and the Caribbean 

Land Degradation Assessment in Drylands 

Land Cover Classification System 

Land Degradation 

Least Developed Countries 

the Landscapes for People, Food and Nature 

Lund-Potsdam-Jena General Ecosystem Simulator 

Long-Range Transboundary Air Pollution 

Topographic Factors 


XXVIII

LU 

MA 

MADRPM 

MAF 

MAFF 

MDBA 

MDGS 

MENARID 

MGAP 

MNP 

MODIS 

NAAS 

NAIP 

NAMA 

NAP 

NAPA 

NBSAP 

NBSS&LUP 

NDVI 

NENA 

NEPAD 

NEST 

NGO 

NISF 

NLWRA 

NOAA 

AVHRR 

NPK 

NPL 

NRC 

NRCAN 

NREL 

NRI 

NRM 

NRSA 

NSW 

NT 

NUE 

Land Use 

Millennium Ecosystem Assessment 

Ministère De l’Agriculture Du Développement Rural et Des Pêches Maritimes 

New Zealand Ministry of Agriculture and Forestry 

Ministry of Agriculture, Forestry and Fishery of Japan 

Murray–Darling Basin Authority (Australia) 

Millennium Development Goals 

Integrated Natural Resources Management in the Middle East And North Africa 

Ministry of Livestock, Agriculture and Fisheries 

Netherlands Environmental Assessment Agency 

Moderate Resolution Imaging Spectroradiometer 

National Academy of Agricultural Sciences of India 

National Agricultural Investment Plan 

Nationally Appropriate Mitigation Action 

(1) National Action Programme; (2) National Action Plan 

National Adaptation Programme of Action 

National Biodiversity Strategy and Action Plan 

National Bureau Of Soil Survey And Land Use Planning 

Normalized Difference Vegetation Index 

Near East And North Africa Region 

The New Partnership for Africa’s Development 

Nigerian Environmental Study Action Team 

Non-Governmental Organization 

National Institute for Soils And Fertilizers 

National Land and Water Resources Audit 

National Oceanic and Atmospheric Administration - Advanced Very High Resolution 

Radiometer 

Nitrogen (N), Phosphorus (P) and Potassium (K) 

National Priorities List 

National Research Council USA 

Natural Resources Canada 

National Resource Ecology Laboratory 

National Resources Inventory Program 

Natural Resources Management 

National Remote Sensing Agency (India) 

New South Wales 

No-Tillage 

Nitrogen Use Efficiency 


XXIX

OECD 

OM 

ÖNORM 

ÖPUL 

ORNL-CDIAC 

OSWER 

PAH 

PAM 

PCB 

PCM 

PEA 

PHC 

PL 

PLAR 

PMID 

PNUD 

POC 

POP 

PVC 

Radar- 

AMEDAS 

RAPA 

RELMA 

ROTAP 

RSN 

RUSLE 

SAGYP-CFA 

SAV 

SCAN 

SCARPS 

SCWMRI 

SD 

SDGS 

SEC 

SEEA 

SEED 

SF 

SFR 

Organization for Economic Co-Operation And Development 

Organic Matter 

National Standard Published By the Austrian Standards Institute 

Austrian Environment Programme for Agriculture 

Oak Ridge National Laboratory-Carbon Dioxide Information Analysis Center 

Office of Solid Waste and Emergency Response 

Polycyclic Aromatic Hydrocarbon 

Polyacrylamide 

Polychlorinated Biphenyl 

Pyrogenic Carbonaceous Matter 

Participatory Expert Assessment 

Petroleum Hydrocarbon 

Plastic Limit 

Participatory Learning-Action-Research 

Pubmed Identifier 

Programa De Las Naciones Unidas Para El Desarrollo 

Particulate Organic Carbon 

Persistent Organic Pollutant 

Polyvinyl Chloride 

Radar-Automated Meteorological Data Acquisition System 

Regional Office for Asia and the Pacific 

Sida’s Regional Land Management Unit 

Review Of Transboundary Air Pollution 

Residual Soil Nitrogen 

Revised Universal Soil Loss Equation 

Secretaría De Agricultura, Ganadería Y Pesca – Consejo Federal Agropecuario 

Submerged Aquatic Vegetation 

Soil Climate Analysis Network 

Salinity Control and Reclamation Projects 

Soil Conservation and Watershed Management Research Institute 

Soil Degradation 

Sustainable Development Goals 

Staff Working Documents of European Commission 

System of Environmental Economic Accounting 

Sustainable Energy and Environment Division 

Subsistence Farming 

Stock-Flow-Resources 


XXX

SKM 

SLAM 

SLC 

SLM 

SMAP 

SMOS 

SOC 

SOE 

SOER 

SOLAW 

SOM 

SOTER 

SOW-VU 

SPARROW 

SPC 

SPI 

SRI 

SSA 

SSM 

SSR 

SSSA 

ST 

STATSGO 2 

STEP-AWBH 

SWC 

SWSR 

TEEB 

TEOM 

TOC 

TOMS 

TOT 

TSBF 

UN 

UNCCD 

UNCED 

UNDCPAC 

UNDESA 

UNDP 

Sinclair Knight Merz 

Sustainable Land and Agro-Ecosystem Management 

Soil Landscapes of Canada 

Sustainable Land Management 

Soil Moisture Active Passive 

Soil Moisture Ocean Salinity 

Soil Organic Carbon 

State of the Environment 

European Environment State and Outlook Report 

State Of Land and Water 

Soil Organic Matter 

Soil and Terrain Database 

Centre for World Food Studies of the University Of Amsterdam 

Spatially Referenced Regressions on Watershed Attributes 

Secretariat of the Pacific Community 

Science-Policy Interface 

Salinity Risk Index 

Sub-Saharan Africa 

Sustainable Soil Management 

Shift Soil Remediation 

Soil Science Society of America 

Soil Taxonomy 

Digital General Soil Map of the United States 

Soil, Topography, Ecology, Parent Material – Atmosphere, Water, Biotic, Human Model 

Soil and Water Conservation 

Status of the World’s Soil Resources 

Economics of Ecosystems and Biodiversity 

Tapered Element Oscillating Microbalances 

Total Organic Carbon 

Total Ozone Mapping Spectrometer 

Transfer of Technology 

Tropical Soil Biology and Fertility 

United Nations 

United Nation Convention to Combat Desertification 

United Nations Conference on Environment And Development 

United Nations Desertification Control Program Activity Center 

United Nations Department of Economic And Social Affairs 

United Nations Development Program 


XXXI

UNEP DEWA 

UNEP 

UNESCO 

UNFCCC 

UNFPA 

UNISDR 

UNSO 

USDA 

USEPA 

USGS 

USLE 

UXO 

WANA 

WCED 

WFP 

WMO 

WOCAT 

WOTR 

WRB 

WRI 

WWF 

United Nations Environment Programme and Department of Early Warning and 

Assessment 

United Nations Environment Programme 

United Nations Educational, Scientific and Cultural Organization 

United Nations Framework Convention on Climate Change 

United Nations Population Fund (Formerly the United Nations Fund for Population 

Activities) 

United Nations Office for Disaster Risk Reduction 

United Nations Development Programme - Office to Combat Desertification and Drought 

United States Department Of Agriculture 

United States Environmental Protection Agency 

United States Geological Survey 

Universal Soil Loss Equation 

Unexploded Ordnance 

West Asia-North Africa 

World Commission on Environment and Development 

United Nations World Food Programme 

World Meteorological Organization 

World Overview of Conservation Approaches and Technologies 

Watershed Organization Trust 

World Reference Base for Soil Resources 

World Resources Institute 

World Wildlife Fund 


XXXII

List of tables 

Table 1.1 | Chronology of introduction of major concepts in pedology and holistic soil management | 7 

Table 1.2 | Ecosystem services provided by the soil and the soil functions that support these services. | 11 

Table 2 | Soil functions related to the water cycle and ecosystem services | 22 

Table 2.2 | Examples of global trends in soil management and their effects on the ecosystem services 

mediated by water. | 24 

Table 3.1 | Generalized ecosystem service rating of specific soil groups (WRB) 7 | 42 

Table 4.1 | Soil carbon lost globally due to land use change over the period 1860 to 2010 (PgC) | 58 

Table 4.2 | Threats to soil resource quality and functioning under agricultural intensification | 64 

Table 4.3 | Artificial areas in Corine Land Cover Legend | 65 

Table 4.4 | Artificial areas in Emilia Romagna according to the Corine Land Cover Legend and sealing index | 66 

Table 5.1 | World population by region | 90 

Table 5.2 | The ten most populous countries 1950, 2013, 2050 and 2100 | 90 

Table 6.1 | Distribution of Soil Organic Carbon Stocks and Density by IPCC Climate Region | 112 

Table 6.3 | Estimate of the historic SOC depletion from principal biomes. Source: Lal, 1999. | 116 

Table 6.4 | Estimates of historic SOC depletion from major soil orders | 117 

Table 6.5 | Estimates of historic SOC loss from accelerated erosion by water and wind | 117 

Table 6.6 | Distribution of salt-affected soils in drylands different continents of the world | 125 

Table 6.7 | Major components of soil nutrient mass balances for N, P and K | 134 

Table 7.1 | Erosion and crop yield reduction estimates from post-2000 review articles | 177 

Table 8.1 | Recent Milestones in soil governance and sustainable development | 227 

Table 8.2 | The 5 Pillars of Action of the Global Soil Partnership. | 227 

Table 9.1 | Characteristics and distribution of agro-ecological zones in Africa | 245 

Table 9.2 | Classes of nutrient loss rate (kg ha-1 yr-1) | 260 

Table 9.4 | Definitions of the five land-cover classes on which the land-cover change study was based | 274 

Table 9.5 | Summary of soil threats status, trends and uncertainties in Africa South of the Sahara. | 277 

Table 10.1 | Soil organic carbon change in selected countries in Asia | 299 

Table 10.2 | Harmonized area statistics of degraded and wastelands of India | 306 

Table 10.3 | Emission factors of drained tropical peatland under different land uses and the 95 percent 

confidential interval | 309 

Table 10.4 | Summary of Soil Threats Status, trends and uncertainties in Asia | 318 

Table 11.1 | The percentage of agricultural land area of total land area in the countries of the European | 332 

Table 11.2 | The areas of saline soils in the countries with major extent of soil salinization in the European 

region | 342 

Table 11.3 Types and extent of soil degradation in Ukraine | 352 

Table 11.4 | Summary of soil threats status, trends and uncertainties in Europe and Eurasia | 358 

Table 12.1 | Summary of Soil Threats Status, trends and uncertainties in in Latin America and the 

Caribbean | 389 


XXXIII

Table 13.1 | Land degradation caused by water erosion in the NENA region (1000 ha) | 404 

Table 13.2 | Soil degradation caused by wind erosion in the NENA region (1000 ha) | 405 

Table 13.3 | Summary of soil threats: Status, trends and uncertainties in the Near East and North Africa | 

432 

Table 14.1 | Summary of soil threats status, trends and uncertainties in North America | 469 

Table 15.1 | Summary of current primary drivers of land-use and the associated implications for soil 

resources in the Southwest Pacific region | 484 

Table 15.2 | Current population, project population (UNDESA, 2013) and Gross Domestic Product per 

capita (World Bank, 2014) for countries of the region. | 484 

Table 15.3 | Estimated annual land–atmosphere (net) carbon (C) exchange rate for New Zealand’s major 

vegetation types | 489 

Table 15.4 | Summary of soil threats status, trends and uncertainties in the Southwest Pacific | 509 

List of boxes 

Box 1.1 | Guidelines for Action | 6 

Box 5.1 | Minefields | 95 

Box 5.2 | Migration/Refugee Camps | 95 

Box 5.3 | Combined effects of war and strife on soils | 95 

Box 6.1 | Livestock-related budgets within village territories in Western Niger | 135 

Box 6.2 | Nutrient balances in urban vegetable production in West African cities | 137 

Box 1 | The catastrophe of the Aral Sea | 354 


XXXIV

List of figures 

Figure 2.1 | Overview of ecosystem processes involved in determining the soil C balance. | 14 

Figure 2.3 | Global (a) nitrogen (N) and (b) phosphorus (P) fertilizer use between 1961 and 2012 split for 

the different continents in Mt P per year. Source: FAO, 2015. | 19 

Figure 2.4 | Applied and excess nitrogen and phosphorus in croplands. Nitrogen and phosphorus 

inputs and excess were calculated using a simple mass balance model, extended to include 175 crops. 

To account for both the rate and spatial extent of croplands, the data are presented as kg per ha of the 

landscape: (a) applied nitrogen, including N deposition; (b) applied phosphorus; (c) excess nitrogen; 

and (d) excess phosphorus. Source: West et al., 2014. | 20 

Figure 3.1 | Nutrient availability in soils. Source: Fischer et al., 2008. | 36 

Figure 3.2 | Global soil rooting conditions. Source: Fischer et al., 2008. | 36 

Figure 3.3 | Soil Moisture storage capacity. Source: Van Engelen, 2012. | 38 

Figure 3.4 | Soil Organic Carbon pool (tonnes C ha -1 ). | 39 

Figure 3.5 | Soil erodibility as characterized by the k factor. Source: Nachtergaele and Petri, 2011. | 40 

Figure 3.6 | Soil workability derived from HWSD. Source: Fischer et al., 2008. | 40 

Figure 3.7 | Soil suitability for cropping at low input, based on the global agro-ecological zones study. 

Source: Fischer et al., 2008. | 41 

Figure 3.8 | GLASOD results. Source: Oldeman, Hakkeling and Sombroek, 1991. | 44 

Figure 3.9 | Example of the effect of land use on indicative factors for ecosystem goods and services | 45 

Figure 3.10 | Soil compaction risk derived from intensity of tractor use in crop land and from livestock 

density in grasslands. Source: Nachtergaele et al., 2011 | 46 

Figure 4.1 | Global Land Cover. Source: Latham et al., 2014. | 51 

Figure 4.2 | Distribution of land cover in different regions. Source: Latham et al., 2014. | 51 

Figure 4.3 | Historical land use change 1000 – 2005. Source: Klein Goldewijk et al., 2011. | 54 

Figure 4.4 | Soil carbon and nitrogen under different land cover types. Source: Smith et al. (in press). | 57 

Figure 4.5 | Maps of change in soil carbon due to land use change and land management from 1860 to 

2010 from three vegetation models. Pink indicates loss of soil carbon, blue indicates carbon gain. The 

models were run with historical land use change. This was compared to a model run with only natural 

vegetation cover to diagnose the difference in soil carbon due to land cover change. Both model runs 

included historical climate and CO 2 

change. Source: Smith et al. (in press). | 58 

Figure 4.6 | Schematic diagram showing areas sealed (B) as a result of infrastructure development for 

a settlement (A). Source: European Union, 2012. | 66 

Figure 4.7 | (A) Panoramic view of Las Medulas opencast gold mine (NW Spain). The Roman extractive 

technique – known as ‘ruina montis’ – involved the massive use of water that resulted in important 

geomorphological changes; (B) Weathered gossan of the Rio Tinto Cu mine, considered the birthplace 

of the Copper and Bronze Ages; (C) typical colour of Rio Tinto (‘red river’ in Spanish), one of the best 

known examples of formation of acid mine waters. These are inhabited by extremophile organisms. 

| 69 


XXXV

Figure 4.8 | Eh-pH conditions of thionic/sulfidic soils and of hyperacid soils. Source: Otero et al., 

2008. | 70 

Figure 4.9 | Use of different Technosols derived from wastes in the recovery of hyperacid soils and 

waters in the restored mine of Touro (Galicia, NW Spain). | 72 

Figure 4.10 | Global distribution of (a) atmospheric S deposition, (b) soil sensitivity to acidification, (c) 

atmospheric N deposition, and (d) soil carbon to nitrogen ratio (soils most sensitive to eutrophication 

have a high C:N ratio; eutrophication is caused by N). Source: Vet et al., 2014; Batjes, 2012; FAO, 

2007. | 75 

Figure 5.1 | Percentage of female landholders around the world. Source: FAO, 2010. | 92 

Figure 5.2 | Major land deals occurring between countries in 2012. Source: Soil Atlas, 2015/Rulli et al., 

2013. | 93 

Figure 6.1 | Spatial variation of soil erosion by water. High rates (>ca. 20 t ha -1 y -1 ) mainly occur on 

cropland in tropical areas. The map gives an indication of current erosion rates and does not assess 

the degradation status of the soils. The map is derived from Van Oost et al., 2007 using a quantile 

classification. | 102 

Figure 6.2 | Location of active and fixed aeolian deposits. Source: Thomas and Wiggs, 2008. | 103 

Figure 6.3 | Soil relict in the Jadan basin, Ecuador. Photo by G. Govers | 103 

In this area overgrazing led to excessive erosion and the soil has been completely stripped from most 

of the landscape in less than 200 years, exposing the highly weathered bedrock below. The person is 

standing on a small patch of the B-horizon of the original soil that has been preserved. Picture credit: 

Gerard Govers. | 103 

Figure 6.4 | Dust storm near Meadow, Texas, USA | 106 

Figure 6.5 | Distribution of carbon in biomass between ORNL-CDIAC Biomass and JRC Carbon 

Biomass Map | 113 

Figure 6.6 | Prevalence of carbon in the topsoil or biomass | 114 

Figure 6.7 | Proportion of carbon in broad vegetation classes for soil and biomass carbon pool | 115 

Figure 6.8 | Estimated dominant topsoil pH. Source: FAO/IIASA/ISRIC/ISS-CAS/JRC, 2009. | 124 

Figure 6.9 | Historical and predicted shift of the urban/rural population ratio. Source: UN, 2008. | 130 

Figure 6.10 | Urbanisation of the best agricultural soils. | 131 

Figure 6.11 | Major components of the soil nutrient balance. 

The red discontinuous line marks the soil volume over which the mass balance is calculated. Green 

arrows correspond to inputs and red arrows to losses. ΔS represents the change in nutrient stock. | 133 

Figure 6.12 | The flows of water and energy through the soil-vegetation horizon | 140 

Figure 6.13 | The soil-water characteristic curve linking matric potential, to the soil’s volumetric water 

content. 

Source: Tuller and Or, 2003. | 141 

Figure 6.14 | The soil’s hydraulic conductivity, K (cm day -1 ) in relation to the matric potential, (MPa). 

As the matric potential becomes more negative the soil’s water content drops (see Figure 6.16) which 

increases the tortuosity and slows the flow of water. Source: Hunter College. 3 | 142 


XXXVI

Figure 6.15 | Factors controlling soil water spatial variability and the scales at which they are 

important. Source: Crow et al., 2010) | 143 

Figure 6.16 | (a) Global distribution of average soil moisture depth in the top 1 m of the soil. (b) 

Seasonal variability in soil moisture calculated as the standard deviation of monthly mean soil 

moisture over the year. (c-d) Global trends (1950-2008) in precipitation and 1 m soil moisture. (e-f) As 

for (c-d) but for 1990-2008. Results for arid regions and permanent ice sheets are not shown. Source: 

Sheffield and Wood, 2007. | 145 

Figure 7.1 | The 11 dimensions of society’s ‘social foundation’ and the nine dimensions of the 

‘environmental ceiling’ of the planet. Source: Vince and Raworth, 2012. | 170 

Figure 7.2 | Conceptual framework for comparing land use and trade-offs of ecosystem services. 

Source: Foley et al., 2005. | 171 

Figure 7.3 | Response curves of mean ecosystem service indicators per 1-km 2 across Great Britain. 

Source: Maskell et al., 2013. | 173 

The curves are fitted using generalized additive models to ordination axes constrained by; (a) 

proportion of intensive land (arable and improved grassland habitats) within each 1-km square from 

CS field survey data; (b) mean long-term annual average rainfall (1978–2005); and (c) mean soil pH 

from five random sampling locations in each 1-km square. All X axes are scaled to the units of each 

constraining variable | 173 

Figure 7.4 | The food wedge and the effect of soil change on the area of the wedge. Source: Keating et 

al., 2014. 

The relative sizes of the effects of soil change on the food wedge are not drawn to scale. | 174 

Figure 7.5 | Direct impacts of soil threats on specific soil functions of relevance to plant production. | 

176 

Figure 7.6 | Some soil-related feedbacks to global climate change to illustrate the complexity and 

potential number of response pathways. Source: Heimann and Reichstein, 2008. | 183 

Figure 7.7 | Definition of soil moisture regimes and corresponding evapotranspiration regimes. 

Source: Seneviratne et al., 2010. 

EF denotes the evaporative fraction, and EFmax its maximal value. | 186 

Figure 7.8 | Estimation of evapotranspiration drivers (moisture and radiation) based on observationdriven 

land surface model simulation. Source: Seneviratne et al., 2010. 

The figure displays yearly correlations of evapotranspiration with global radiation Rg and precipitation 

P in simulations from the 2nd phase of the Global Soil Wetness Project (GSWP, Dirmeyer et al., 2006) 

using a two-dimensional color map, based on Teuling et al. 2009, redrawn for the whole globe. 

(Seneviratne et al., 2010) | 187 

Figure 7.9 | A conceptual sketch of how vulnerability, exposure and external events (climate, weather, 

geophysical) contribute to the risk of a natural hazard. Source: IPCC, 2012. | 196 

Figure 7.10 | Trends in landslide frequency and mortality on Asia. Source: FAO, 2011; EM-DAT, 2010. | 197 

Figure 9.1 | Agro-ecological zones in Africa South of the Sahara. Source: Otte and Chilonda, 2002. | 245 

Figure 9.2 | Extent of urban areas and Urbanization Indexes for the Sub-Saharan African countries. 

Source: Schneider, Friedl and Potere, 2010. | 253 


XXXVII

Figure 9.3 | The fertility rate (the number of children a woman is expected to bear during her lifetime) 

for 1970 and 2005. Source: Fooddesert.org | 255 

Figure 9.4 | Percentage of population living below the poverty line. Source: CIA World Factbook, 

2012. | 255 

Table 9.3 | Estimated nutrient balance in some SSA countries in 1982-84 and forecasts for 2000. Surce: 

Stoorvogel and Smaling, 1990; Roy et al., 2003. | 262 

Figure 9.5 | Major land use systems in Senegal. Source: FAO, 2010. | 264 

Figure 9.6 Proportional extent of major land use systems in the Senegal. Source: Ndiaye and Dieng, 

2013. | 264 

Figure 9.7 | Extent of dominant degradation type in Senegal. Source: FAO, 2010. | 265 

Figure 9.8 | Average rate of degradation in Senegal. Source: FAO, 2010. | 265 

Figure 9.9 | Impact of degradation on ecosystem services in the local study areas in Senegal. Source: 

Ndiaye and Dieng, 2013. | 266 

Figure 9.10 | Broad soil patterns of South Africa. Source: Land Type Survey Staff, 2003. | 268 

Figure 9.11 | The national stratification used for land degradation assessment in South Africa, 

incorporating local municipality boundaries with 18 land use classes. Source: Pretorius, 2009. | 270 

Figure 9.12 | Actual water erosion prediction map of South Africa. Source: Le Roux et al., 2012. | 271 

Figure 9.13 | Topsoil pH derived from undisturbed (natural) soils. Source: Beukes, Stronkhorst and 

Jezile, 2008a. | 273 

Figure 9.14 | Change in land-cover between 1994 and 2005 as part of the Five Class Land-cover of 

South Africa after logical corrections. Source: Schoeman et al., 2010. | 275 

Figure 10.1 | Length of the available growing period in Asia (in days yr -1 ). Source: Fischer et al., 2012. | 289 

Figure 10.2 | Threats to soils in the Asia region by country. | 291 

Figure 10.3 | Nitrogen surplus or depletion, and nutrient use efficiency in crop production in Asia and 

the Middle East in 2010. | 304 

Figure 10.4 | Degradation and wastelands map of India. Source: ICAR and NAAS, 2010. | 305 

Figure 10.5 | Indonesian peatland map overlaid with land cover map as of 2011. Source: Wahyunto et 

al., 2014. | 310 

Figure 10.6 | Distribution map of radioactive Cs concentration in soil in Fukushima prefecture 

(reference date of 5 November, 2011). Source: Takata et al., 2014. | 312 

Figure 10.7 | Distribution map of the parameters of USLE and classification of estimated soil loss. 

Class I: less than 1 tonnes ha -1 yr -1 ; Class II: 1-5 tonnes ha -1 yr -1 ; Class III: 5-10 tonnes ha -1 yr -1 ; Class IV: 10-30 

tonnes ha -1 yr -1 ; Class V: 30-50 tonnes ha -1 yr -1 ; Class VI: more than 50 tonnes ha -1 yr -1 . Source: Kohyama 

et al., 2012. | 313 

Figure 10.8 | Estimate CH 4 emission from rice paddy in Asia. Source: Yan et al., 2009. | 315 

Figure 11.1 | Terrestrial eco-regions of the European region. Source: Olson et al., 2001. | 333 

Figure 11.2 | Soil salinization on the territory of the European region. Source: Afonin et al., 2008; Toth 

et al., 2008; GDRS, 1987. | 343 

Figure 11.3 | Some types and extent of soil degradation in Ukraine. Source: Medvedev, 2012. | 353 


XXXVIII

Figure 11.4 | Soil map and soil degradation extent in Uzbekistan. Source: Arabov, 2010. | 355 

Figure 12.1 | Biomes in Latin America and the Caribbean. Source: Olson et al., 2001. | 367 

Figure 12.2 | Extent of the urban area and the urbanization index for Latin American and Caribbean 

countries. | 374 

Figure 12.3 | shows soil organic carbon contents and stocks (taking into account soil bulk density) 

in different Mexican ecosystems. Carbon concentrations (left) and carbon stocks (right) in the main 

ecosystems of Mexico. In both cases the bars with the strongest tone indicate a primary forest, closed 

pasture or permanent agriculture. Bars with the softer tone indicate a secondary forest, open pasture 

or annual agriculture. Source: Cruz-Gaistardo, 2014. | 376 

Figure 12.4 | Organic carbon stock (or density) in soils of Latin America and the Caribbean, expressed 

in Gigagrams per hectare. Source: Gardi et al., 2014. | 378 

Figure 12.5 | Tree cover in the tonne 2000 and forest loss in the period 2000-2014. (A) Brazil, centered 

at 5.3°S, 50.2°W; (B) Mexico and Guatemala, centered at 16.3°N, 90.8°W and (C) Perú, centered at 

8.7°S, 74.9°W; (D) Argentina, centered at 27.0°S, 62.3°W and (E) Chile, centered at 72.5°S, 37.4°W. Source: 

Hansen et al., 2013. | 379 

Figure 12.6 | Expansion of the agricultural frontier under rainfed conditions in the north of Argentina. 

Source: Viglizzo & Jobbagy, 2010. | 383 

Figure 12.7 | Percentage of areas affected by wind (a) and water erosion (b) in Argentina. Source: 

Prego et al., 1988. | 385 

Figure 12.8 | Predominant types of land degradation in Cuba. Source: FAO, 2010. | 387 

Figure 12.9 | Extent of land degradation in land use system units in Cuba. Source: FAO, 2010. | 387 

Figure 12.10 | Intensity of land degradation in Cuba. Source: FAO, 2010. | 388 

Figure 13.1 | Land use systems in the Near East and North Africa. Source: FAO, 2010. | 403 

Figure 13.2 | Extent of the urban areas and Urbanization Indexes for the Near East and North African 

countries. Source: Schneider, Friedl and Potere, 2009. | 410 

Figure 13.3 | Layout of the project site source (a) and conceptual design and layout of bioremediation 

system (b). Source: Balba et al., 1998. | 421 

Figure 13.4 | Rate of water erosion in Iran. Source: Soil Conservation and Watershed Management 

Research Institute. | 424 

Figure 13.5 | Shows days with dust storms in 2012, while Figure 13.6 shows the origin of dust storms in 2012. | 425 

Figure 13.6 | Internal and external dust sources in recent years in Iran. Source: University of Tehran, 2013. | 426 

Figure 13.7 | Assessment of Water (a) and Wind Erosion (b) in Tunisia | 427 

Figure 13.8 | Soil Conservation in Tunisia | 428 

Figure 13.9 | Type of ecosystem service most affected. | 429 

Figure 14.1 | Level II Ecological regions of North America. Source: Commission for Environmental 

Cooperation, 1997. | 446 

Figure 14.2 | Map of Superfund sites in the contiguous United States Yellow indicates final EPA 

National Priorities List sites and red indicates proposed sites. Source: EPA, 2014a. | 449 

Figure 14.3 | Areas in United States threatened by salinization and sodification. Source: NRCS 1 | 451 


XXXIX

Figure 14.4 | Risk of soil salinization in Canada 2011. Source: Clearwater et al., 2015. | 452 

Figure 14.5 | Risk of water erosion in Canada 2011. Source: Clearwater et al., 2015. | 461 

Figure 14.6 | Risk of wind erosion in Canada 2011. Source: Clearwater et al., 2015. | 462 

Figure 14.7 | Soil organic carbon change in Canada 201. Source: Clearwater et al., 2015. | 463 

Figure 14.8 | Residual soil N in Canada 2011. Source: Clearwater et al., 2015. | 465 

Figure 14.9 | Indicator of risk of water contamination by phosphorus (IROWC-P) in Canada in 2011. 

Source: Clearwater et al., 2015. | 466 

Figure 15.1 | Nations in the Southwest Pacific region and the extent of Melanesia, Micronesia and 

Polynesian cultures. Figure based on base map imagery: exclusive economic zone boundaries (EEZ)v 8 

2014, Natural Earth 11 3.2.0 | 478 

Figure 15.2 | Change in the percentage area of all land prepared for crops and pastures under different 

tillage practices in Australia, 1996-2010 Source: SOE, 2011. | 486 

Figure 15.3 | (a) Trends in winter rainfall in south-western Australia for the period 1900–2012. Source: 

Australian Bureau of Meteorology 1 . 

The 15-year running average is shown by the black line. (b) Annual mean temperature anomaly time 

series map for south-western Australia (1910–2012), using a baseline annual temperature (1961–1990) 

of 16.3 °C. The 15-year running average is shown by the black line. | 501 

Figure 15.4 | Percentage of sites sampled (2005–12) with soil pH at 0–10 cm depth below the 

established target of pHCa 5.5 (left) and the critical pHCa 5.0 (right). Grey indicates native vegetation 

and reserves. Source: Gazey, Andrew and Griffin, 2013. | 502 

Figure 15.5 | Agricultural lime sales 2005–12 in the south-west of Western Australia based on data for 

85–90 percent of the market. | 503 

Figure 15.6 | MODIS image for 0000 23 September 2009 showing Red Dawn extending from south 

of Sydney to the Queensland/NSW border and the PM 10 concentrations measured using Tapered 

Element Oscillating Microbalances (TEOM) at the same time at ground stations. | 506 

Figure A 1 | (a) A Histosol profile and (b) a peatbog in East-European tundra. | 529 

Figure A 2 | (a) An Anthrosol (Plaggen) profile and (b) associated landscape in the Netherlands. | 531 

Figure A 3 | (a) A Technosol profile and (b) artefacts found in Technosol. | 533 

Figure A 4 | (a) A Cryosol profile and (b) associated landscape in West Siberia, Yamal Peninsula. | 535 

Figure A 5 | (a) A Leptosol profile in the Northern Ural Mountains and (b) associated landscape. | 537 

Figure A 6 | Vertisol gilgai patterns and associated soils: (a) linear gilgai pattern located on a 

moderately sloping hillside in western South Dakota. Distance between repeating gilgai cycle is about 

4 m. (b) Normal gilgai pattern occurring on a nearly level clayey terrace near College Station, TX. After a 

rainfall event microlows have been partially filled with runoff water from microhighs - repeating gilgai 

cycle about 4 m in linear length. (c) Trench exposure of soils excavated across normal gilgai pattern - 

repeating gilgai cycle about 4 m in linear length. Dark-colored deep soil in microlow (leached A and Bss 

horizons) with light-colored shallow calcareous soils associated with diaper in microhigh (Bssk and Ck 

horizons). The diaper has been thrust along oblique slickenside planes towards soils surface. Vertical 

depth of soil trench in about 2 m. (d) Close up of dark-colored soil associated with microlow and light 


XL

colored diaper associated with microhigh of the trench in (c). | 539 

Figure A 7 | (a) A Solonetz profile and (b) the associated landscape in Hungary. | 541 

Figure A 8 | (a) A Solonchak profile and (b) a salt crust with halophytes. | 543 

Figure A 9 | (a) A Podzol profile and (b) an associated landscape, West-Siberian Plain. | 545 

Figure A 10 | (a) A giant Podzol profile and (b) an associated landscape, Brazil. | 546 

Figure A 11 | (a) A Ferralsol profile and (b) an associated landscape, Brazil. | 548 

Figure A 12 | (a) A Nitisol profile and (b) the associated landscape with termite mounds, Brazil. | 550 

Figure A 13 | (a) A Plinthosol profile, (b) details of the plinthic horizon and (c) the associated landscape, 

South Africa. | 552 

Figure A 14 | (a) A Planosol profile and (b) the associated landscape, Argentina. | 554 

Figure A 15 | (a) A Gleysol profile and (b) associated landscape in the East European tundra. | 556 

Figure A 16 | (a) A Stagnosol profile, (b) stagnic color patterns, (c) marble-like horizontal surface and 

(d) an associated landscape. | 558 

Figure A 17 | (a) An Andosol profile and (b) the associated landscape in Japan. | 560 

Figure A 18 | (a) A Chernozem profile (Photo by J. Deckers) and (b) the associated landscape in the 

Central Russian Uplands. | 562 

Figure A 19 | (a) A Kastanozem profile and (b) the associated landscape in Mongolia. | 564 

Figure A 20 | (a) A Phaeozem profile and (b) the associated landscape, Argentinian Pampa. | 566 

Figure A 21 | (a) An Umbrisol profile, (b) associated vegetation and (c) an associated landscape. | 568 

Figure A 22 | (a) A Durisol profile and (b) the associated landscape, Ecuador. | 570 

Figure A 23 | (a) A Calcisol profile, (b) an associated landscape and (c and d) secondary carbonates in 

Calcisols. | 572 

Figure A 24 | (a) A Gypsisol profile and (b) an associated landscape. | 574 

Figure A 25 | (a) A Retisol profile, (b) the “retic” pattern in a Retisol and (c) the associated landscape, 

Belgium. | 576 

Figure A 26 | (a) An Acrisol profile and (b) the associated landform in Kalimantan, Indonesia. | 578 

Figure A 27 | (a) A Lixisol profile and (b) the associated landscape, Brazil. | 580 

Figure A 28 | (a) An Alisol profile and (b) the associated landscape, Belgium. | 582 

Figure A 29 | (a) A Luvisol profile and (b) the associated landscape, China. | 584 

Figure A 30 | (a) A Cambisol profile and (b) the associated landscape, China. | 586 

Figure A 31 | (a) A Regosol profile and (b) the associated landscape, China. | 588 

Figure A 32 | (a) An Arenosol profile in South Korea and (b) an Arenosol profile in New Mexico. | 590 

Figure A 33 | (a) A Fluvisol profile in Wisconsin and (b) a Fluvisol profile in Germany. | 592 

Figure A 34 | (a) A Wassent profile and (b) the associated landscape, the Netherlands. | 594 

Figure A 35 | Global Soil Map of the World based on HWSD and FAO Revised Legend (Nachtergaele 

and Petri, 2008) | 595 


XLI


XLII

Preface 

The main objectives of The State of the World’s Soil Resources are: (a) to provide a global scientific 

assessment of current and projected soil conditions built on regional data analysis and expertise; (b) to 

explore the implications of these soil conditions for food security, climate change, water quality and quantity, 

biodiversity, and human health and wellbeing; and (c) to conclude with a series of recommendations for action 

by policymakers and other stakeholders. 

The book is divided into two parts. The first part deals with global soil issues (Chapters 1 to 8). This is 

followed by a more specific assessment of regional soil change, covering in turn Africa South of the Sahara, 

Asia, Europe, Latin America and the Caribbean, the Near East and North Africa, North America, the Southwest 

Pacific and Antarctica. (Chapters 9 to 16). The technical and executive summaries are published separately. 

In Chapter 1 the principles of the World Soil Charter are discussed, including guidelines for stakeholders to 

ensure that soils are managed sustainably and that degraded soils are rehabilitated or restored. For long, soil 

was considered almost exclusively in the context of food production. However, with the increasing impact of 

humans on the environment, the connections between soil and broader environmental concerns have been 

made and new and innovative ways of relating soils to people have begun to emerge in the past two decades. 

Societal issues such as food security, sustainability, climate change, carbon sequestration, greenhouse gas 

emissions, and degradation through erosion and loss of organic matter and nutrients are all closely related 

to the soil resource. These ecosystem services provided by the soil and the soil functions that support these 

services are central to the discussion in the report. 

In Chapter 2 synergies and trade-offs are reviewed, together with the role of soils in supporting ecosystem 

services, and their role in underpinning natural capital. The discussion then covers knowledge - and knowledge 

gaps - on the role of soils in the carbon, nitrogen and water cycles, and on the role of soils as a habitat for 

organisms and as a genetic pool. This is followed in Chapter 3 by an overview of the diversity of global soil 

resources and of the way they have been assessed in the past. Chapter 4 reviews the various anthropogenic 

and natural pressures - in particular, land use and soil management – which cause chemical, physical and 

biological variations in soils and the consequent changes in environmental services assured by those soils. 

Land use and soil management are in turn largely determined by socio-economic conditions. These 

conditions are the subject of Chapter 5, which discusses in particular the role of population dynamics, market 

access, education and cultural values as well as the wealth or poverty of the land users. Climate change and 

its anticipated effects on soils are also discussed in this chapter. 

Chapter 6 discusses the current global status and trends of the major soil processes threatening ecosystem 

services. These include soil erosion, soil organic carbon loss, soil contamination, soil acidification, soil 

salinization, soil biodiversity loss, soil surface effects, soil nutrient status, soil compaction and soil moisture 

conditions. 

Chapter 7 undertakes an assessment of the ways in which soil change is likely to impact on soil functions 

and the likely consequences for ecosystem service delivery. Each subsection in this chapter outlines key soil 

processes involved with the delivery of goods and services and how these are changing. The subsections 

then review how these changes affect soil function and the soil’s contribution to ecosystem service delivery. 

The discussion is organized according to the reporting categories of the Millennium Ecosystem Assessment, 

including provisioning, supporting, regulating and cultural services. 

Status of the World’s Soil Resources | Main Report Preface 

1

Chapter 8 of the report explores policy, institutional and land use management options and responses to 

soil changes that are available to governments and land users. 

The regional assessments in Chapters 9 to 16 follow a standard outline: after a brief description of the main 

biophysical features of each region, the status and trends of each major soil threat are discussed. Each chapter 

ends with one or more national case studies of soil change and a table summarizing the results, including the 

status and trends of soil changes in the region and related uncertainties. 

Status of the World’s Soil Resources | Main Report Preface 

2

Global soil resources 

Coordinating Lead Authors: 

Maria Gerasimova (Russia), Thomas Reinsch (United States), Pete Smith (United Kingdom) 

Contributing Authors: 

Lucia Anjos (Brazil), Susumu Asakawa (Japan), Ochirbat Batkhishig (Mongolia), James 

Bockheim (United States), Robert Brinkman (Netherlands), Gabrielle Broll (Germany), Mercedes 

Bustamante (Brazil), Marta Camps Arbestain (ITPS/New Zealand), Przemyslaw Charzynski (Poland), 

Joanna Clark (United Kingdom), Francesca Cotrufo (United States), Maur’cio Rizzato Coelho (Brazil), 

Jane Elliott (Canada), Maria Gerasimova (Russia), Robert I. Griffiths (United Kingdom), 

Richard Harper (Australia), Jo House (United Kingdom), Peter Kuikman (Netherlands), 

Tapan Kumar Adhya (India), Richard McDowell (New Zealand), Freddy Nachtergaele (Belgium), 

Masami Nanzyo (Japan), Christian Omutu (Kenya), Genxing Pan (China), Keith Paustian (United States), 

Dan Pennock (ITPS/Canada), Cornelia Rumpel (France), Jaroslava Sobocká (Slovakia), 

Mark Stolt (United States), Mabel Susna Pasos (Argentina), Charles Tarnocai (Canada), Tibor Toth 

(Hungary), Ronald Vargas (Bolivia), Paul West (United States), Larry P. Wilding (United States), 

Ganlin Zhang (ITPS/China), Juan JosŽ Ibánez (Spain), Felipe Macias (Spain). 

Reviewing Authors: 

Dominique Arrouays (ITPS/France), Richard Bardgett (United Kingdom), Marta Camps Arbestain 

(ITPS/New Zealand), Tandra Fraser (Canada), Ciro Gardi (Italy), Neil McKenzie (ITPS/Australia), 

Luca Montanarella (ITPS/EC), Dan Pennock (ITPS/Canada) and Diana Wall (United States). 

Status of the World’s Soil Resources | Main Report Global soil resources 

3

1 | Introduction 

1.1 | The World Soil Charter 

“Soils are fundamental to life on earth.” 

We know more about soil than ever before, yet perhaps a smaller percentage of people than at any point 

in human history would understand the truth of this statement. The proportion of human labour devoted to 

working the soil has steadily decreased through the past century, and hence the experience of direct contact 

with the soil has lessened in most regions. Soil is very different in this regard from food, energy, water and air, 

to which each of us requires constant and secure access. Yet human society as a whole depends more than 

ever before on products from the soil as well as on the more intangible services it provides for maintenance of 

the biosphere. 

Our goal in this report is to make clear these essential connections between human well-being and the soil, 

and to provide a benchmark against which our collective progress to conserve this essential resource can be 

measured. 

The statement that begins this section is drawn from the opening sentence of the preamble of the revised 

World Soil Charter (FAO, 2015): 

Soils are fundamental to life on Earth but human pressures on soil resources are reaching 

critical limits. Careful soil management is one essential element of sustainable agriculture and 

also provides a valuable lever for climate regulation and a pathway for safeguarding ecosystem 

services and biodiversity. 

The World Soil Charter presents a series of nine principles that summarize our current understanding of the 

soil, the multi-faceted role it plays, and the threats to its ability to continue to serve these roles. As such, the 

nine principles form a succinct and comprehensive introduction to this report. 

Principles from the World Soil Charter: 

Principle 1: Soils are a key enabling resource, central to the creation of a host of goods and services integral 

to ecosystems and human well-being. The maintenance or enhancement of global soil resources is essential if 

humanity’s overarching need for food, water, and energy security is to be met in accordance with the sovereign 

rights of each state over their natural resources. In particular, the projected increases in food, fibre, and fuel 

production required to achieve food and energy security will place increased pressure on the soil. 

Principle 2: Soils result from complex actions and interactions of processes in time and space and hence 

are themselves diverse in form and properties and the level of ecosystems services they provide. Good soil 

governance requires that these differing soil capabilities be understood and that land use that respects the 

range of capabilities be encouraged with a view to eradicating poverty and achieving food security. 

Principle 3: Soil management is sustainable if the supporting, provisioning, regulating, and cultural services 

provided by soil are maintained or enhanced without significantly impairing either the soil functions that 

enable those services or biodiversity. 


4

The balance between the supporting and provisioning services for plant production and the regulating 

services the soil provides for water quality and availability and for atmospheric greenhouse gas composition 

is a particular concern. 

Principle 4: The implementation of soil management decisions is typically made locally and occurs within 

widely differing socio-economic contexts. The development of specific measures appropriate for adoption by 

local decision-makers often requires multi-level, interdisciplinary initiatives by many stakeholders. A strong 

commitment to including local and indigenous knowledge is critical. 

Principle 5: The specific functions provided by a soil are governed, in large part, by the suite of chemical, 

biological, and physical properties present in that soil. Knowledge of the actual state of those properties, their 

role in soil functions, and the effect of change – both natural and human-induced – on them is essential to 

achieve sustainability. 

Principle 6: Soils are a key reservoir of global biodiversity, which ranges from micro-organisms to flora and 

fauna. This biodiversity has a fundamental role in supporting soil functions and therefore ecosystem goods 

and services associated with soils. Therefore it is necessary to maintain soil biodiversity to safeguard these 

functions. 

Principle 7: All soils – whether actively managed or not – provide ecosystem services relevant to global 

climate regulation and multi-scale water regulation. Land use conversion can reduce these global commongood 

services provided by soils. The impact of local or regional land-use conversions can be reliably evaluated 

only in the context of global evaluations of the contribution of soils to essential ecosystem services. 

Principle 8: Soil degradation inherently reduces or eliminates soil functions and their ability to support 

ecosystem services essential for human well-being. Minimizing or eliminating significant soil degradation 

is essential to maintain the services provided by all soils and is substantially more cost-effective than 

rehabilitating soils after degradation has occurred. 

Principle 9: Soils that have experienced degradation can, in some cases, have their core functions and their 

contributions to ecosystem services restored through the application of appropriate rehabilitation techniques. 

This increases the area available for the provision of services without necessitating land use conversion. 

These nine principles lead to guidelines for action by society (Box 1.1). The guidelines are introduced with 

a clear statement of our collective goal: ‘The overarching goal for all parties is to ensure that soils are managed 

sustainably and that degraded soils are rehabilitated or restored.’ This opening statement is followed by a series of 

specific guidelines for different segments of human society. Future updates of this report will document our 

success in implementation of these guidelines, and in achieving the goal set by the signatories of the World 

Soil Charter. 


5

Box 1.1 | Guidelines for Action 

(from the World Soil Charter) 

The overarching goal for all parties is to ensure 

that soils are managed sustainably and that 

degraded soils are rehabilitated or restored. 

Good soil governance requires that actions 

at all levels – from states, and, to the extent 

that they are able, other public authorities, 

international organizations, individuals, groups, 

and corporations – be informed by the principles 

of sustainable soil management and contribute 

to the achievement of a land-degradation neutral 

world in the context of sustainable development. 

All actors and, specifically, each of the following 

stakeholder groups are encouraged to consider the 

following actions: 

Actions by Individuals and the Private Sector 

1. All individuals using or managing soil must 

act as stewards of the soil to ensure that 

this essential natural resource is managed 

sustainably to safeguard it for future 

generations. 

2. Undertake sustainable soil management in the 

production of goods and services. 

Actions by Groups and the Science Community 

1. Disseminate information and knowledge on 

soils. 

2. Emphasize the importance of sustainable 

soil management to avoid impairing key soil 

functions. 

Actions by Governments 

1. Promote sustainable soil management that is 

relevant to the range of soils present and the 

needs of the country. 

2. Strive to create socio-economic and 

institutional conditions favourable to 

sustainable soil management by removal 

of obstacles. Ways and means should be 

pursued to overcome obstacles to the 

adoption of sustainable soil management 

associated with land tenure, the rights 

of users, access to financial services and 

educational programmes. Reference is made 

to the Voluntary Guidelines on the Responsible 

Governance of Tenure of Land, Forests and 

Fisheries in the Context of National Food 

Security adopted by the Committee on World 

Food Security in May 2012. 

3. Participate in the development of multi-level, 

interdisciplinary educational and capacitybuilding 

initiatives that promote the adoption 

of sustainable soil management by land users. 

4. Support research programs that will provide 

sound scientific backing for development 

and implementation of sustainable soil 

management relevant to end users. 

5. Incorporate the principles and practices 

of sustainable soil management into 

policy guidance and legislation at all levels 

of government, ideally leading to the 

development of a national soil policy. 

6. Explicitly consider the role of soil management 

practices in planning for adaptation to and 

mitigation of climate change and maintaining 

biodiversity. 

7. Establish and implement regulations to limit 

the accumulation of contaminants beyond 

established levels to safeguard human health 

and wellbeing and facilitate remediation of 

contaminated soils that exceed these levels 

where they pose a threat to humans, plants, 

and animals. 

8. Develop and maintain a national soil 

information system and contribute to the 

development of a global soil information 

system. 

9. Develop a national institutional framework for 

monitoring implementation of sustainable soil 

management and overall state of soil resources. 

Actions by International Organizations 

10. Facilitate the compilation and dissemination 

of authoritative reports on the state of the 

global soil resources and sustainable soil 

management protocols. 

11. Coordinate efforts to develop an accurate, 

high-resolution global soil information system 

and ensure its integration with other global 

earth observing systems. 

12. Assist governments, on request, to establish 

appropriate legislation, institutions, 

and processes to enable them to mount, 

implement, and monitor appropriate 

sustainable soil management practices. 


6

1.2 | Basic concepts 

Prior to the 20th century, soil was considered almost exclusively in the context of agriculture and food 

production. As the global impact of humanity on natural resources has increased over the past 150 years, the 

connections between soil and broader environmental concerns began to be made. The recognition of these 

connections has accelerated through time, and new and innovative ways of relating soils to people have begun 

to emerge in past the two decades. The rise in complexity of soil knowledge and application was synthesized 

by Bockheim et al. (2005) (Table 1.1) in their summary of milestones in pedology; concepts introduced since 

2005 have been added by the authors of this chapter. We can see that the number and breadth of concepts 

have been expanding rapidly over the past two decades. 

Period Pedology Soil management 

Pre -1 880 

Concept of soil as a medium for plant growth and as a 

weathered rock layer. 

1880–1900 

Appearance of fundamental pedology concepts: soil as a natural 

body; soil horizons/profiles; soil-forming factors; early ideas of 

soil geography. 

1900–1940 

Global acceptance of concepts of soil as a natural body and soilforming 

factors; development of first regional soil classification 

systems; soil surveys initiated; identification of key soil-forming 

processes. 

Soil conservation 

1940– 1960 

Factors of soil formation and genesis of soils clarified; 

development of global soil taxonomic systems; intensified soil 

mapping. 

1960–1985 

Refinement of global soil taxonomic systems; identification 

of pedon concept; development of early soil models and soil 

cover pattern concept; recognition of co-evolution of soils and 

landforms. 

World Soil Charter 

(1981) Land capability/ 

suitability assessment 

Assessment of humaninduced 

degradation 

(GLASOD) 

1985– 2000 

Increased understanding of soil processes; refinement of 

global soil models; further refinement of global soil taxonomic 

systems; development of statistical and computer-based soil 

information systems. 

Sustainable soil 

management 

Soil quality 

Soil health 

2000-2015 

Earth System Science, pedosphere, digital soil mapping, 

pedodiversity, ethnopedology, pedometrics, proximal sensing, 

soil systems. hydropedology, critical zone. 

Soil security 

Carbon sequestration 

Table 1.1. Chronology of introduction of major concepts in pedology and holistic soil management 

(after Bockheim et al., 2005). 


7

The connections between soils and societal issues – such as food security, sustainability, climate change, 

carbon sequestration, greenhouse gas emissions, and degradation through erosion and loss of organic 

matter and nutrients – are central to the recently developed concept of soil security (McBratney, Field and 

Koch, 2014). Soil security has been defined as the maintenance or improvement of the world’s soil resources 

so that they can provide sufficient food, fibre, and fresh water, contribute to energy sustainability and climate 

stability, maintain biodiversity, and deliver overall environmental protection and ecosystem services (Bouma 

and McBratney, 2013). 

There have been major developments over the past three decades in our broader understanding of human 

impact on the earth and of frameworks to assess this impact. The structure and content of this report comprise 

a synthesis of themes and concepts from many major initiatives in environmental science and pedology. The 

most important of these themes and concepts are discussed in the following paragraphs. 

Sustainable soil management 

The concept of sustainable development is most closely associated with the 1987 report of the United 

Nations World Commission on Environment and Development, better known as the Brundtland Commission 

after its chairperson, Gro Harlem Brundtland of Norway (World Commission on Environment and 

Development, 1987). The report popularized a compelling definition of sustainability: development that meets 

the needs of the present without compromising the ability of future generations to meet their own needs. 

The concept of sustainability has since been widely applied to many aspects of human society, including wide 

application in soil science and land management generally. As defined in the World Soil Charter, sustainable 

soil management comprises activities that maintain or enhance the supporting, provisioning, regulating 

and cultural services provided by soils without significantly impairing either the soil functions that enable 

those services or biodiversity. The concept of sustainable soil management is central to pillar one of the Global 

Soil Partnership 1 : “Promote sustainable management of soil resources for soil protection, conservation, and 

sustainable production”. 

Soil degradation and threats to soil functions 

The concept of soil degradation and its assessment have been developed as part of more holistic assessments 

of human-induced degradation carried out by FAO, UNEP and other UN agencies. 

An early initiative was the Global Assessment of Soil Degradation (GLASOD) project undertaken in the late 

1980s to inventory soil degradation. GLASOD evaluated 13 types of soil degradation: water erosion (topsoil 

loss and mass movement, including rill and gully formation), wind erosion (topsoil loss, terrain deformation 

– primarily dune activity), and overblowing (surface burial from aeolian deposition), loss of nutrients and/ 

or organic matter, salinization, acidification, pollution, compaction and physical degradation, waterlogging, 

and subsidence of organic soils. GLASOD has not been updated (see Chapter 3 for more details). 

The Soil Thematic Strategy of the European Union (CEC, 2006) formalized the concept of threats to soil and 

its many functions. Five specific threats are identified under Article 6 of the draft Soil Framework Directive 

proposed in the Strategy: (1) erosion by wind and water; (2) organic matter decline; (3) compaction; (4) 

salinization; and (5) landslides of soil and rock material. Elsewhere in the proposed Directive, soil sealing (‘the 

permanent covering of the soil with an impermeable surface’ p.15) and soil contamination (‘the intentional or 

unintentional introduction of dangerous substances on or in the soil’ p. 18) are also identified as threats. 

1 The Global Soil Partnership was initiated by FAO and the EU in 2011. For a description of the five pillars, see Table 8.1 in Chapter 8. For a full description of 

the Partnership, see www.fao.org/globalsoilpartnership). 


8

Soil functions and ecosystem services 

The assessments of threats to soil functions leads to a need to formally identify the functions that the soil 

performs. The proposed Soil Framework Directive (CEC, 2006) of the European Union recognizes seven soil 

functions that are vulnerable to soil threats: 

1. biomass production, including agriculture and forestry 

2. storing, filtering and transforming nutrients, substances and water 

3. biodiversity pool, such as habitats, species and genes 

4. physical and cultural environment for humans and human activities 

5. source of raw materials 

6. acting as a carbon pool 

7. archive of geological and archaeological heritage. 

The EU Soil Thematic Strategy was developed at the same time as the Millennium Ecosystem Assessment 

(MA, 2005) initiated by the United Nations in 2000. The goal of the MA was to assess the consequences 

of ecosystem change for human well-being and to lay the scientific basis for actions that would promote 

conservation and sustainable use of ecosystems. The MA was built on the framework for ecosystem services 

developed by Daily, Matson and Vitousek (1997) and Costanza et al. (1997). 

The categories of ecosystem services were formalized by the Millennium Ecosystem Assessment into four 

broad classes: provisioning, regulating, supporting, and cultural services. The range of major ecosystem 

services provided by soil, and the specific soil functions that enable those services, are summarized in Table 1.2. 

Soils and natural capital 

The services provided by soils are primarily determined by the three core soil properties (texture, mineralogy, 

and organic matter), which together form the natural capital of soils (Palm et al. 2007). Soil texture and 

mineralogy are inherent properties of soil that are initially inherited from the parent materials and which 

change only very slowly over time. In a natural state, soil organic matter (SOM) reaches equilibrium with the 

environment in which the soil forms, but SOM responds quickly to human-induced changes. Management 

of SOM is central to sustainable soil management because of its rapid response to change and our ability to 

manipulate it. 

Planetary boundaries and safe operating space for humanity 

Specific soil processes are central to Earth-system processes that provide the safe operating space for 

humanity – the concept of ‘planetary boundaries’ that cannot be exceeded without causing potentially 

disastrous consequences for humanity (Röckstrom et al. 2009; Steffen et al., 2015). Currently stresses in the 

nitrogen cycle, climate change, and biodiversity loss are suggested to be beyond safe operating boundaries. 

Human impact on the natural reservoir of soil biodiversity and on the rate of N and C cycling in soils is a 

significant aspect of this stress. Whereas GLASOD had highlighted nutrient depletion through crop production 

without the application of sufficient manure and fertilizer to replenish nutrient loss, the concept of planetary 

boundaries also focuses our attention on over-application of nutrients in some regions and its consequences 

for atmospheric and hydrological systems. Addressing the nutrient deficit in regions such as Sub-Saharan 

Africa while remaining within the safe operating space for humanity requires a significant reduction of 

nutrient additions in area of excess inputs (Steffen et al., 2015). 


9

Biodiversity 

Biodiversity cuts across most of the concepts presented above, and loss of biodiversity is identified by 

Röckstrom et al. (2009) as one of three components currently operating beyond safe planetary boundaries. 

Biodiversity is more than simply an ecosystem service, even though specific benefits can be identified from the 

biodiversity pool. This cross-cutting importance of biodiversity was formalized in the Convention on Biological 

Diversity signed in 1992 at the United Nations Conference on Environment and Development in Brazil. Soils are 

widely recognized as a major reservoir of global biodiversity, and preservation of this (largely unknown) pool 

of biodiversity is essential. 

Ecosystem service 

Soil functions 

Supporting services: Services that are necessary for the production of all other ecosystem services; their 

impacts on people are often indirect or occur over a very long time 

Soil formation 

∑ Weathering of primary minerals and 

release of nutrients 

∑ Transformation and accumulation of 

organic matter 

∑ Creation of structures (aggregates, 

horizons) for gas and water flow and root 

growth 

∑ Creation of charged surfaces for ion 

retention and exchange 

Primary production 

∑ Medium for seed germination and root 

growth 

∑ Supply of nutrients and water for plants 

Nutrient cycling 

∑ Transformation of organic materials by soil 

organisms 

∑ Retention and release of nutrients on 

charged surfaces 

Regulating services: benefits obtained from the regulation of ecosystem processes 

Water quality regulation 

∑ Filtering and buffering of substances in soil 

water 

∑ Transformation of contaminants 

Water supply regulation 

∑ Regulation of water infiltration into soil 

and water flow within the soil 

∑ Drainage of excess water out of soil and 

into groundwater and surface water 


10

Climate regulation 

∑ Regulation of CO 2 

, N 2 

O, and CH 4 

emissions 

Erosion regulation 

∑ Retention of soil on the land surface 

Provisioning Services: products (‘goods’) obtained from ecosystems of direct benefit to people 

Food supply 

∑ Providing water, nutrients, and physical 

support for growth of plants for human 

and animal consumption 

Water supply 

∑ Retention and purification of water 

Fibre and fuel supply 

∑ Providing water, nutrients, and physical 

support for growth of plant growth for 

bioenergy and fibre 

Raw earth material supply 

∑ Provision of topsoil, aggregates, peat etc. 

Surface stability 

∑ Supporting human habitations and related 

infrastructure 

Refugia 

∑ Providing habitat for soil animals, birds 

etc. 

Genetic resources 

∑ Source of unique biological materials 

Cultural services: nonmaterial benefits which people obtain from ecosystems through spiritual 

enrichment, aesthetic experiences, heritage preservation and recreation 

Aesthetic and spiritual 

∑ Preservation of natural and cultural 

landscape diversity 

∑ Source of pigments and dyes 

Heritage 

∑ Preservation of archaeological records 

Table 1.2: Ecosystem services provided by the soil and the soil functions that support these services. 


11

The contribution of many of the concepts outlined above is apparent throughout the World Soil Charter. 

This synthesis of concepts is perhaps most evident in the definition of sustainable soil management used in 

the World Soil Charter: 

Soil management is sustainable if the supporting, provisioning, regulating and cultural services 

provided by soil are maintained or enhanced without significantly impairing either the soil 

functions that enable those services or biodiversity. 

The concepts of soil functions, the threats to functions, and the ecosystem services provided by soils are 

central both to the structure of this book and to the content of each chapter. 

References 

Bockheim, J.G., Gennadiyev, A.N., Hammer, R.D. & Tandarich, J.P. 2005. Historical development of key 

concepts in pedology. Geoderma, 124: 23-36. 

Bouma, J. & McBratney, A.B. 2013. Framing soils as an actor when dealing with wicked environmental 

problems. Geoderma, 200: 130 -1 39. 

Commission of the European Communities (CEC). 2006. Communication from the Commission to the 

Council, the European Parliament, the European Economic and Social Committee and the Committee of the 

Regions. Thematic Strategy for Soil Protection. COM 231 Final, Brussels. 

Costanza R., d’Arge R., de Groot R., Farber S., Grasso M., Hannon B., Limburg, K., Naeem, S., O’Neill, 

R.V., Paruelo, J., Raskin, R.G., Sutton, P. & van den Belt, M. 1997. The value of the world’s ecosystem services 

and natural capital. Nature, 387: 253-260. 

Daily, G.C., P.A. Matson & Vitousek P.M. 1997. Ecosystem services supplied by soil. In G. Daily, ed. pp. 113 -1 32. 

Nature’s Services: Societal Dependence on Natural Ecosystems. Washington, DC, Island Press. 412 pp. 

FAO. 2015. World Soil Charter (also available at http://www.fao.org/3/a-mn442e.pdf) 

Hole, F.D. & Campbell, J.B. 1985. Soil Landscape Analysis. London, Routledge & Kegan Paul. 196 pp. 

McBratney, A. B., Field, D. J., & Koch, A. 2014. The dimensions of soil security. Geoderma, 213: 203-213. 

Millennium Ecosystem Assessment. 2005. Ecosystems and Human Well-Being: Synthesis. Washington, 

DC, Island Press. 800 pp. 

NRC 1998. The Canadian System of Soil Classification. 3 rd ed. Canada, Ottawa. 187 pp. 

Palm, C., Sanchez, P., Ahamed, S. & Awiti, A. 2007. Soils: a contemporary perspective. Annu. Rev. Environ. 

Resour., 32: 99 -1 29. 

Rockström, J., Steffen, W., Noone, K., Persson, Å., Chapin III, F.S., Lambin, E.F., Lenton, T.M., Scheffer, 

M., Folke, C., Schellnhuber, H.J., Nykvist, B., de Wit, C.A., Hughes, T., van der Leeuw, S., Rodhe, H., Sörlin, 

S., Snyder, P.K., Costanza, R., Svedin, U., Falkenmark, M., Karlberg, L., Corell, R.W., Fabry, V.J., Hansen, J., 

Walker, B., Liverman, D., Richardson, K., Crutzen, P. & Foley, J.A. 2009. A safe operating space for humanity. 

Nature, 461: 472-475. 

Steffen, W., Richardson, K., Rockström, J., Cornell, S.E., Fetzer, I., Bennett, E.M., Biggs, R. , Carpenter, 

S.R., de Vries, W., de Wit, C.A., Folke, C., Gerten, D., Heinke, J., Mace, G.M., Persson, L. M., Ramanathan, 

V., Reyers, B., & Sörlin, S. 2015. Planetary boundaries: Guiding human development on a changing planet. 

Science, 347 (6223): 1259855. 10 pp. 

World Commission on Environment and Development. 1987. Report of the World Commission on Environment 

and Development: Our Common Future. UK, Oxford, Oxford University Press. 383 pp. 


12

2 | The role of soils 

in ecosystem processes 

Soils play a critical role in delivering ecosystem services. Management to change an ecosystem process in 

support of one regulating ecosystem service can either provide co-benefits to other services or require tradeoffs 

(Robinson et al., 2013; Dominati, Patterson, and Mackay, 2010). Recent reviews have provided examples 

of some of these synergies and trade-offs (Smith et al., 2013) and illustrated the role of soils in supporting 

ecosystem services and underpinning natural capital (Robinson, Lebron and Vereecken, 2009, Robinson et al., 

2014, Dominati, Patterson and Mackay, 2010). In this chapter, we present current knowledge – and knowledge 

gaps – on the role of soils in the carbon, nitrogen and water cycles, and on their role as a habitat for organisms 

and as a genetic pool. 

2.1 | Soils and the carbon cycle 

Carbon (C) storage is an important ecosystem function of soils that has gained increasing attention in 

recent years due to its interactions with the earth’s climate system. Soil is a major C reservoir that holds more 

carbon than is contained in the atmosphere and terrestrial vegetation combined. All three of these reservoirs 

are in constant exchange. In many soils, soil organic matter (SOM), which contains roughly 55–60 percent 

C by mass, comprises most or all of the C stock – referred to as soil organic carbon (SOC). In arid and semiarid 

soils, significant inorganic C (IC) can be present as pedogenic carbonate minerals or ‘caliche’ (typically Ca/ 

MgCO 3 

), formed from the reaction of biocarbonate (derived from CO 2 

in the soil) with free base cations, which 

can then be precipitated in subsoil layers (Nordt, Wilding and Drees, 2000). Also soils derived from carbonatecontaining 

parent material (e.g. limestone) can have significant amounts of inorganic carbon. However, in 

most cases changes in inorganic C stocks are slow and not amenable to traditional soil management practices. 

Hence inorganic carbon does not play a significant role in terms of management of ecosystem services. For 

this reason, the further discussion of soil C in this chapter will focus on soil organic carbon. 

A general overview of the ecosystem C cycle as it interacts with soils is given in Figure 2.1. The major input of 

organic C to soils is provided by the uptake and fixation of CO 2 

by plants (the net result of photosynthesis and 

above- and below-ground plant respiration), and by the subsequent incorporation of plant residue C (both 

above- and below-ground) into soil. Some of the fixed plant C may be removed by harvest before entering the 

soil. Conversely, C additions from offsite sources (e.g. compost, manure) may occur. Organic matter on and 

in the soil is subject to comminution and mixing by soil fauna and to enzymatic breakdown and metabolism 

by microorganisms, resulting in release of CO 2 

via microbial respiration (also referred to as organic carbon 

mineralization). Microbial transformations as well as interactions of organic matter with soil minerals greatly 

influence the stabilization of organic C and its rate of mineralization. In flooded soils, emissions of methane 

(CH 4 

) from microbial metabolism can represent a significant gaseous C efflux. Erosion can also directly 

Status of the World’s Soil Resources | Main Report The role of soils in ecosystem processes 

13

affect the soil C balance through the removal and/or deposition of the C contained in the transported soil. 

Leaching of dissolved organic (DOC) and dissolved inorganic carbon (DIC) through the soil profile and out into 

groundwater and surface water represents an additional loss pathway that can be significant in some soils. 

Maintaining and increasing SOC stocks through improved land use and management practices can help 

to counteract increasing atmospheric CO 2 

concentrations (Paustian et al., 1998, Smith et al., 2007; Whitmore, 

Kirk and Rawlins, 2014). Increasing soil C content also improves other chemical and physical soil properties, 

such as nutrient storage, water holding capacity, aggregation and sorption of organic and/or inorganic 

pollutants (Kibblewhite, Ritz and Swift, 2008). Carbon sequestration in soils may therefore be a cost-effective 

and environmentally friendly way to store C. It can also enhance other ecosystem services derived from soil, 

such as agricultural production, clean water supply, and biodiversity by increasing SOM content and thereby 

improving soil quality (Lal, 2004). 

O site OM 

additions 

Plant 

photosynthesis 

CO 2 

CO 2 

Plant 

respiration 

Harvest 

Soil deposition 

Litter 

fall 

SOM 

turnover and 

stabilization 

Decomposition & 

microbial 

respiration 

CO 2 

Soil erosion 

Drawing by A. Swan and K. Paustian 

DOC & DIC 

Figure 2.1 Overview of ecosystem processes involved in determining the soil C balance. 

2.1.1 | Quantitative amounts of organic C stored in soil 

Organic C stocks in the world’s soils have been estimated to comprise 1 500 Pg of C down to 1 m depth and 

2 500 Pg down to 2 m (Batjes, 1996). Recent studies, based on newer estimates for the C stored in boreal 

soils under permafrost conditions, suggest that soil C storage may be even greater, accounting for as much as 

2000 Pg to 1 m depth (Tarnocai et al., 2009). Although the highest C concentrations are found in the top 30 

cm of soil, the major proportion of total C stock in many soils is present below 30 cm depth (Batjes, 1996). In 

the northern circumpolar permafrost region, at least 61 percent of the total soil C is stored below 30 cm depth 

(Tarnocai et al., 2009). 


14

2.1.2 | Nature and formation of soil organic C 

Soil organic matter (SOM) is composed of plant litter compounds as well as of microbial decomposition 

products. SOM is thus a complex biogeochemical mixture derived from organic material in all stages of 

decomposition (von Lützow et al., 2006; Paul, 2014). Due to microbial degradation and mineralization to 

CO 2 

(and CH 4 

in anaerobic environments), the majority of plant litter compounds added to soil remain for a 

relatively short time (from a few days to a few years). This is particularly the case if the organic compounds 

are added on the soil surface. However, some organic matter compounds may persist in the soil for decades or 

centuries or even for millennia (Paul et al., 1997; von Lützow et al., 2006). It is increasingly accepted that, despite 

their recalcitrant nature, plant litter compounds (e.g. lignin) themselves do not substantially contribute to 

SOM persistence in soil (Thévenot, Dignac and Rumpel, 2010). Longer term stabilization is generally conferred 

through interactions with soil minerals (e.g. through surface binding or occlusion within microaggregates), 

which reduce SOM exposure to enzymatic degradation (Sollins, Homann and Caldwell, 1996; Six, Elliott and 

Paustian, 2000; Schmidt et al., 2011). Thus, the location of SOM within the soil matrix has a much stronger 

influence on its turnover than its chemical composition (Chabbi, Kögel-Knabner and Rumpel, 2009; Dungait 

et al., 2012) 

One consequence of the role of reactive mineral surfaces in SOC stabilization is that the surface area of 

the soil mineral fraction, which is finite and a function of soil texture (e.g. clay, silt or sand content) and of 

mineralogy, may set an upper limit for the amount of SOM that a particular soil can hold (Six, Elliott and 

Paustian, 2002). A recent conceptual model (Figure 2.2) by Cotrufo et al. (2013), based on studies showing 

that microbially-derived decomposition products make up most of the mineral-stabilized organic matter, 

postulates that relatively labile litter compounds with higher microbial growth yield efficiency contribute 

proportionally more to the stable mineral-associated SOM pool than do more recalcitrant plant compounds 

with low microbial growth yield efficiency. This concept is in agreement with the current understanding that 

microbial material is building up much of the stabilised SOM pool (Miltner et al., 2012). 

Figure 2.2 

Conceptual model of 

interactions between litter 

quality, microbial products 

and soil mineral interactions 

affecting the formation 

and stabilization of organic 

matter. 

Source: Cotrufo et al., 2013. 


15

With this emphasis on the importance of SOM location within the soil, microbial accessibility of organic 

material at very small scales has become a focus of research in recent years (Lehmann, Kinyangi and Solomon, 

2007). The development of powerful new tools like X-ray spectroscopy and secondary ion mass spectrometry 

now allows the visualisation of organo-mineral interactions at nanoscale. As a result, the location and 

distribution of organic matter within the soil mineral matrix may now be assessed in more detail (Lehmann et 

al., 2008). However, before the results obtained with these tools yield information concerning soil C formation 

at macroscale, upscaling and integration of spatial heterogeneity is necessary (Mueller et al., 2013). 

As well as exhibiting tremendous heterogeneity in terms of its composition, the distribution of SOC within 

the soil is also very heterogeneous, particularly with respect to depth within the soil profile. Whereas upper 

soil layers receive greater amounts of aboveground litter (‘shoot C’ from leaves and stems), subsoil C originates 

primarily from root-derived C as well as from plant- and microbial-derived dissolved organic carbon (DOC) 

transported down the soil profile. Root C has a greater likelihood of being preserved in soil compared to shoot 

C (Balesdant and Balabane, 1996) and studies suggest that root C therefore accounts for a larger proportion 

of SOM (Rasse, Rumpel and Dignac, 2005). In general, C cycling and C formation is most active in topsoil 

horizons, whereas stabilised C with longer turnover times makes up a greater proportion of the total SOC 

found in deep soil horizons (Scharpenseel and Becker-Heidmann, 1989; Trumbore, 2009). The accumulation 

of stabilised C with long residence times in deep soil horizons may be due to continuous transport, temporary 

immobilisation and microbial processing of DOC within the soil profile (Kalbitz and Kaiser, 2012) and/or 

efficient stabilisation of root-derived organic matter within the soil matrix (Rasse, Rumpel and Dignac, 2005). 

An additional long-term C pool in many soils is pyrogenic carbonaceous matter, formed from partially 

carbonised (e.g., pyrolysed) biomass during wildfires (Schmidt and Noack, 2000). A portion of this material 

has a highly condensed aromatic chemical structure (often referred to as pyrogenic carbon or black carbon) 

that resists microbial degradation and can persist in soils for long periods (Lehmann et al., 2015). 

2.1.3 | Soil C pools 

For modelling purposes, soil C is usually divided into a number of pools (typically from two to five) in order 

to represent the heterogeneity in residence time of the vast mixture of different organic compounds in soil 

(Smith et al., 1997). A useful three pool split of soil C (excluding litter) – into a labile pool, an intermediate pool, 

and a refractory (stable) pool – is employed in several soil C models, including the Century model (Parton et 

al., 1987). The labile pool represents easily degradable plant material, microbial biomass and labile metabolites, 

and may turn over within a few months or years. Conceptually, the intermediate pool comprises microbiallyprocessed 

organic matter that is partially stabilized on mineral surfaces and/or protected within aggregates, 

with turnover times in the range of decades. The refractory pool, including highly stabilized organic mattermineral 

complexes and pyrogenic C, may remain in soils for centuries or millennia. 

Individual model pools (as opposed to the total C stock) are typically not defined as measureable pools per 

se. The kinetics of the model conceptual pools are instead inferred from C dating and tracer studies, laboratory 

incubations and total SOC dynamics in long-term field experiments (McGill, 1996; Paustian, 1994). Many 

carbon cycle, ecosystem and crop growth models successfully employ this type of functional representation of 

SOM (Krull, Baldock and Skjemstad, 2003; Stockman et al., 2013). Nonetheless, ways to reconcile ‘measurable’ 

and ‘modelable’ pools have been under discussion for a number of years (Elliott, Paustian and Frey, 1996; Smith 

et al., 2002; Dungait et al., 2012). This reconciliation remains a desirable goal for improving understanding of 

SOC dynamics (Schmidt et al., 2011). 


16

2.1.4 | Factors influencing soil C storage 

Fundamentally, the amount of SOC stored in a given soil is determined by the balance of C entering the 

soil, mainly via plant residues and exudates, and C leaving the soil through mineralization (as CO 2 

), driven 

by microbial processes, and to a lesser extent leaching out of the soil as DOC. Locally C can also be lost or 

gained through soil erosion or deposition (Figure 2.1), leading to a redistribution of soil C at local, landscape 

and regional scales. 

Consequently, a main control on SOC storage is the amount and type of residues that are produced by 

plants as the primary producers in the ecosystem. Plant productivity and subsequent senescence and death 

lead, through plant necromass breakdown, to the input of organic C to the soil system. Thus, broadly speaking 

for a given pedoclimatic condition, higher levels of plant residue inputs will tend to support higher SOC stocks, 

and vice versa. C levels of many soils are also influenced by fertiliser additions, which are indispensable for 

sustaining plant productivity in agricultural systems. 

In addition to productivity and plant C inputs, climatic factors, such as soil temperature and water content 

greatly influence soil C storage through their effect on microbial activity. In general, higher soil temperatures 

increase microbial decomposition of organic matter. Temperature is, therefore, taken as major control of 

SOM storage in soil C cycle models, although the temperature sensitivity of decomposition for different SOM 

fractions remains an area of uncertainty (Conant et al., 2011). 

Water also influences soil C storage through several processes. Moist but well-aerated soils are optimal 

for microbial activity. Decomposition rates consequently decrease as soils become drier. However, flooded 

soils have lower rates of organic matter decay due to restricted aeration (e.g. O 2 

depletion due to limited 

O 2 

diffusion in water) and thus may often yield soils with very high amounts of soil C (e.g. peat and muck 

soils). High precipitation may also lead to C transport down the soil profile as dissolved and/or particulate 

organic matter. During extreme events, such as drought, SOM decomposition may initially decrease but may 

subsequently increase after rewetting (Borken and Matzner, 2008). Fire may decrease soil C storage at first, 

but over the longer term may increase C storage through positive effects on plant growth and through input 

of very stable pyrogenic C (Knicker, 2007). 

The quantity and composition of SOC in mineral soils is also strongly dependent on soil type, with clay 

content influencing not only the amount but also the composition of soil C. In clay rich soils, higher organic 

matter content and a higher concentration of O-alkyl C derived from polysaccharides may be expected, 

compared to sandy soils which are characterised by lower C contents and high concentrations of alkyl C 

(Rumpel and Kögel-Knabner, 2011). Aliphatic material may contribute to the hydrophobicity of soils, which 

could lead to reduced microbial accessibility and therefore increased C storage. 

Bioturbation (the reworking of soils by animals or plants) may further influence the amount as well as 

the chemical nature of soil C. It may greatly influence the heterogeneity of soils by creating hotspots. On 

biologically active sites, incorporation and transformation of organic compounds into soil is usually enhanced 

by bioturbation, leading to organo-mineral interactions and increase of C storage (Wilkinson, Richards and 

Humphreys, 2009). 

Microbial decomposition of SOM may be stimulated (or reduced) by labile organic matter input through 

the ‘priming effect’ (Jenkinson, 1971; Kuzyakov, 2002). Positive priming refers to mineralisation of otherwise 

stable C through shifts in microbial community composition (Fontaine, Mariotti and Abbadie et al., 2003). 

However, in some cases, the addition of organic matter to soil may also cause changes in the soil microbial 

communities with regard to the preferentially degraded substrate and therefore impede mineralisation of 

native SOM (Sparling, Cheschire and Mundie, 1982; Kuzyakov, 2002). 

Plant communities are main controlling factors of these processes because they influence organic matter 

input and microbial activity by their effects on soil water, labile C input, pH and nutrient cycling. 


17

2.1.5 | Carbon cycle: knowledge gaps and research needs 

Substantial progress has been made in recent years towards a deeper understanding of the processes 

controlling soil C storage. There has been progress also in improving and deploying predictive models of soil 

C dynamics that can guide decision makers and inform policy. However, it is equally true that many new (and 

some old) gaps in our knowledge have been identified and the need for further research has been assessed. 

Recent research on soil C dynamics has been driven in part by increasing awareness of: (1) the importance 

of small scale variability for microbial C turnover (Vogel et al., 2014); (2) interactions between the C cycle and 

other biogeochemical cycles (Gärdenäs et al., 2011); and (3) the importance of soil C not only at the field scale 

but at regional to global scales (Todd-Brown et al., 2013). 

The most cited knowledge gaps and research needs include: 

Basic understanding 

• Controls on microbial efficiency of organic matter processing, including biodiversity 

• The degree of association or separation of organic matter and microbial decomposer communities in 

the mineral soil matrix 

• Role of soil fauna in controlling carbon storage and cycling 

• Dynamics of dissolved organic carbon and its role in determining C storage and decomposition 

• Pyrogenic C stabilization and interactions of pyrogenic C with native soil C and mineral nutrients 

• Role of soil erosion in the global C cycle 

Predictive modelling and assessment 

• Reconciliation of measured and modelled SOM fractions 

• More explicit representation of microbial controls 

• Improved modelling of C in subsurface soil layers 

• Distributed soil C observational and monitoring networks for model validation 

• More realistic and spatially-resolved representation of soil C in global-scale models 

2.1.6 | Concluding remarks 

Both biotic and abiotic factors control soil C content and dynamics through their effect on plant litter inputs 

and microbial decomposer communities. The understanding of the C cycle and the role of soils as a sink or source 

of CO 2 

depends on our ability to integrate knowledge of physical, chemical and biological processes operating 

at small scales (nm, µm, soil profile) and of the spatial heterogeneity of SOM distribution and decomposition 

processes at increasing scales (field, region, globe). At the global scale, soils are a major component of the 

planet’s C cycle and can have a strong influence on the concentration of CO 2 

in the atmosphere. Thus, land 

management needs to be based on an understanding of the controls on SOM distribution, stabilisation and 

turnover in order to safeguard and increase the organic matter content of our soils. This will be an important 

contribution to both food security and the mitigation of greenhouse gases. 

2.2 | Soils and the nutrient cycle 

Soils support plant growth and so are vital to humanity. They provide nutrients such as nitrogen (N), 

phosphorous (P), potassium (K), Calcium (Ca), Magnesium (Mg), Sulphur (S) and many trace elements that 

support biomass production. Biomass is important for food supply, for energy and fibre production and as a 

(future) source for the chemical industry. Since the 1950s, higher biomass production and yield increases have 

been supported through mineral/synthetic fertilization (Figure 2.3). However, intensification of agricultural 

practices and of land use has in many regions resulted in a decline in the content of organic matter content 

in agricultural soils. In some areas, extensive use of mineral fertilizers has resulted in atmospheric pollution, 

greenhouse gas emissions (e.g. CO 2 

and N 2 

O), water eutrophication and human health risks (Galloway et al., 

2008). 


18

In coming years, human population and demand for food, feed and energy will continue to rise. In order to 

sustain biomass production in the future and to mitigate negative environmental impacts, fertile soils need 

to be preserved. Where soil fertility has declined, it needs to be restored by maintaining sufficient amounts 

of organic matter in soils (Janzen, 2006). This can be achieved by measures of sustainable management (see 

Chapter 8 of this volume), including by targeted additions of mineral and organic amendments to soils. 

The soil function ‘fertility’ refers to the ability of soil to support and sustain plant growth, including through 

making N, P and other nutrients available for plant uptake. This process is facilitated by: (i) nutrient storage 

in soil organic matter; (ii) nutrient recycling from organic to plant-available mineral forms; and (iii) physical 

and chemical processes that control nutrient sorption, availability, displacement and eventual losses to the 

atmosphere and water. 

Managed soils represent a highly dynamic system and it is this very dynamism that makes soils function 

and supply ecosystem services. Overall, the fertility and functioning of soils depend on interactions between 

the soil mineral matrix, plants and microbes. These are responsible for both building and decomposing SOM 

and therefore for the preservation and availability of nutrients in soils. To sustain soil functions, the balanced 

cycling of nutrients in soils must be maintained. 

After carbon (Section 2.1), N is the most abundant nutrient in all forms of life, since it is contained in proteins, 

nucleic acids and other compounds. Humans and animals ultimately acquire their N from plants, which in 

terrestrial ecosystems occurs mostly in mineral form (e.g. NH 4 

+ and NO 3 

‐ ) in soils. The parent material of 

soils does not contain significant amounts of N (as opposed to P and other nutrients). New N enters the soil 

through the fixation of atmospheric N 2 

by a specialized group of soil biota. However, the largest flux of N in 

Figure 2.3 Global (a) nitrogen (N) and (b) phosphorus (P) fertilizer use between 1961 and 2012 split for the different continents in Mt P 

per year. Source: FAO, 2015. 

soils is generated through the continuous recycling of N internal to the plant-soil system: soil mineral N is 

taken up by the plant, it is fixed into biomass, and eventually N returns in the form of plant debris to the soil. 

Here soil biota decompose it, mineralizing part of the N and making it newly available for plant growth, while 

transforming the other part into SOM, which ultimately is the largest stock of stable N in soil. Nitrogen is lost 

from the soil to the water system by leaching and to the atmosphere by gas efflux (NH 3 

, N 2 

O and N 2 

). 

In most natural ecosystems, N availability is a limiting factor to productivity and N cycles tightly in the 

system with minimal losses. Through the cultivation of N 2 

fixing crops, the production and application of 

synthetic N fertilizer, and the deposition of atmospheric N, humans have applied twice as much reactive N 

to soils as the N introduced by natural processes, thereby significantly increasing biomass production on land 

(Vitousek and Matson, 1993). However, since mineral fertilizer use efficiency is generally low and far more 


19

fertilizer is often used than plants actually need, a high percentage of N fertilizer is lost from the soil. This is 

generating a myriad of deleterious cascade effects on the environment and on human health (Galloway et al., 

2008). This phenomenon is spread over most of the globe. However, in some regions of the world, in particular 

Sub-Saharan Africa, which are characterized by eroded soils and where economic constraints limit the use of 

fertilizers, productivity is still strongly constrained by low levels of soil-available N and other nutrients, notably 

P (Figure 2.2). 

Phosphorus is an essential element for all living organisms. It cycles internally in the plant-soil system, moving 

from the parent material through weathering to biochemical molecules (e.g. nucleic acid, phospholipids) and 

back to mineral forms after decomposition (e.g. H 3 

PO 4 

). In natural soils P is among the most limiting nutrients, 

since it is present in small amounts and only available in its soluble forms, which promptly react with calcium, 

iron and aluminum cations to precipitate as highly insoluble compounds. Adsorbed on those compounds, P 

can be lost from soils, entering the aquatic system through erosion and surface runoff. To correct this lack of 

available P, ‘primary’ P is mined and added to soils in the form of mineral fertilizer. This external input has led to 

positive agronomic P balances (McDonald et al., 2011). There are, however, large variations in the world, with 

large surpluses in the United States, Europe and Asia, and deficits in Russia, Africa and South-America (Figure 

2.4). Additionally, since plant P uptake is a relatively inefficient process with roughly 60 percent of the total P 

input to soils not taken up, it has been estimated that the amount of P exported from terrestrial to aquatic 

systems has tripled, with significant impacts on the environment (Bennett, Carpenter and Caraco, 2001). 

Figure 2.4 Applied and excess nitrogen and phosphorus in croplands. Nitrogen and phosphorus inputs and excess were calculated 

using a simple mass balance model, extended to include 175 crops. To account for both the rate and spatial extent of croplands, the 

data are presented as kg per ha of the landscape: (a) applied nitrogen, including N deposition; (b) applied phosphorus; (c) excess 

nitrogen; and (d) excess phosphorus. Source: West et al., 2014. 


20

Management practices need to be implemented that sustain, restore or increase soil fertility and biomass 

production while limiting associated negative impacts. This can be achieved by promoting the accrual of soil 

organic matter and nutrient recycling, applying balanced C amendments and fertilization of N, P and other 

nutrients to meet plant and soil requirements, while limiting overuse of fertilizer. Carbon, N and P cycling in 

soils is coupled by tight stoichiometric relationships (e.g. relatively fixed C:N:P in plants and microorganisms). 

This means that an enduring increase or decrease of carbon in soils cannot be achieved without a proportional 

change in nitrogen and phosphorous (and several other nutrients). This is a fundamental consideration in any 

programs for carbon sequestration and land restoration because of the significant costs. Therefore, their 

management needs to be planned in concert. 

Nutrient management has been extensively studied, with the aim of identifying and proposing management 

practices (e.g. precision agriculture) that improve nutrient use efficiency and productivity while reducing 

potentially harmful losses to the environment (van Groenigen et al., 2010; Venterea, Maharjan and Dolan, 

2011). However, our ability to predict the ecosystem response to balanced fertilization is still limited and the 

relationship requires continued monitoring. Further benefits are anticipated from improved plant varieties 

with root morphologies that have better capacity to extract P from soils or use it more efficiently. 

More generally, further research is needed into organic matter responses to agricultural C inputs and 

into the potential for restoring and increasing soil organic matter to promote long term soil fertility (e.g. 

Lugato, Berti and Giardini, 2006). Hence, we stress the importance of an integrated approach to nutrient 

management which supports plant productivity while preserving or enhancing soil organic matter stocks and 

reducing nutrient losses to the atmosphere or aquatic systems. Prediction and optimization of performance 

would benefit from continued data acquisition across the whole range of climate and environmental and 

agro-ecological conditions. 

2.2.1 | The nutrient cycle: knowledge gaps and research needs 

In the second half of the 20th century, higher biomass yields were supported by higher use of fertilizer 

(N, P) inputs. This is now considered unsustainable in many situations. Alternatives are required that make 

better use of inherent soil fertility, improve resource use efficiency, and prevent losses of N and P. Examples in 

agriculture include sustainable intensification and new crop varieties that have root systems with improved 

extraction capability or which have higher internal P use efficiency. At the food system level, more effective 

nutrient management would benefit from a focus on a ‘5R strategy’: (1) realign P and N inputs; (2) reduce P 

and N losses to water, thereby minimizing eutrophication impacts; (3) recycle the P and N in bio-resources; (4) 

recover P and N from wastes to use as fertilizer; and (5) redefine use and use-efficiency of P in the food chain 

(Withers et al., 2015). 

In addition, a better understanding of biogeochemical processes at the molecular level is needed. This 

should include: (i) research into the role of plant symbionts on the weathering of minerals and support of 

nutrient uptake, and (ii) development of target-specific ‘smart’ agrochemical agents that enhance nutrient 

uptake. 

2.3 | Soils and the water cycle 

Soils provide important ecosystem services through their function within the water cycle. These services include 

provisioning services of food and water security, regulating services associated with moderation and purification of water 

flows, and cultural services such as landscapes and water bodies that meet recreation and aesthetic values (Dymond, 

2014). Water stored in soil is used for the evapotranspiration and plant growth that supply food and fibre. Soil water also 

stabilizes the land surface to prevent erosion and regulates nutrient and contaminant flow. At a catchment and basin 

scale, the capacity of the soil to infiltrate water attenuates stream and river flows and can prevent flooding, while water 

that percolates through soil can replenish groundwater and related streamflow and surface water ecosystems. 


21

The soil functions of accepting, storing, transmitting and cleaning of water shown in Table 2.1 are interrelated. 

Soil water storage depends on the rate of infiltration into the soil and on soil hydraulic conductivity 

that redistributes water within and through the soil profile. Similarly, infiltration and hydraulic conductivity 

are dependent on the water stored in the soil. The initially high rate of infiltration into dry soil declines as 

the soil water content increases and water replaces air in the pore space. Conversely, hydraulic conductivity 

increases with soil moisture content as a greater proportion of the pores are transmitting water. Water 

content and transmission times are also important to the filtering function of soil because contact with soil 

surfaces and residence time in soil are controls on contaminant supply and removal. 

Optimum growth of most plants occurs when roots can access both oxygen and water in the soil. The soil 

must therefore infiltrate water, drain quickly when saturated to allow air to reach plant roots, and retain and 

redistribute water for plant use. The ideal soil for plant production depends on climatic conditions and on the 

soil requirements of the crop. For instance, in dry regions it can be an advantage to have soils with a high clay 

content to retain water, while sandier soils that drain quickly are better suited to wetter regions. 

Soil structural stability and porosity are also important for the infiltration of water into soil. Organic matter 

improves soil aggregate stability. While plant growth and surface mulches can help protect the soil surface, 

a stable, well-aggregated soil structure that resists surface sealing and continues to infiltrate water during 

intense rainfall events will decrease the potential for downstream flooding. Porosity determines the capacity 

of the soil to retain water and controls transmission of water through the soil. In addition to total porosity, the 

continuity and structure of the pore network are important to these functions and also to the further function 

of filtering out contaminants in flow. 

Soil Function Mechanism Consequence Ecosystem service 

Stores 

(Storage) 

Water held in soil pores 

supports plant and 

microbial communities 

Biomass production 

Surface protection 

Food 

Aesthetics 

Erosion control 

Accepts 

(Sorptivity) 

Incident water 

infiltrates into soil with 

excess lost as runoff 

Storm runoff reduction 

Erosion control 

Flood protection 

Transmits 

(Hydraulic conductivity) 

Water entering the soil 

is redistributed and 

excess is transmitted as 

deep percolation 

Percolation to 

groundwater 

Groundwater recharge 

Stream flow 

maintenance 

Cleans 

(Filtering) 

Water passing through 

the soil matrix interacts 

with soil particles and 

biota 

Contaminants 

removed by biological 

degradation/retention 

on sorption sites 

Water quality 

Table 2. Soil functions related to the water cycle and ecosystem services 


22

Another important role of soil water is its support of biota that can degrade compounds into beneficial 

forms that may also retain nutrients. Soil organic matter is important to this role - together with mineral 

soil (especially the clay fraction), SOM provides sorption sites, but sorption capacity is finite. Flow through 

macropores that bypass the soil matrix where biota and sorption sites are generally located can quickly 

transmit water and contaminants through the soil to groundwater or artificial drains. However, for filtering 

purposes a longer, slower route through the soil matrix is more effective. 

Soil management alters the ecosystem services provided by water (Table 2.2). Soil conservation practices 

and sustainable management help to retain regulating ecosystem services such as soil organic matter and 

structural stability. Similarly, the promotion of soil as a C-sink to offset greenhouse gas emissions helps to 

maintain or improve soil functions. On the other hand, deforestation, overgrazing and excessive tillage of 

fragile lands lead to deterioration of the soil structure and to loss of soil function and surface water quality 

(Steinfield et al., 2006). 

Anthropogenic modifications to the water cycle can aid soil function. In dry regimes, inadequate soil 

moisture can be mitigated through supplementary irrigation, and where excessive precipitation causes 

problems, waterlogging can be relieved by land drainage. However, irrigation and drainage can have 

consequences for water regulation services. Irrigation that enables a shift to intensive land use can increase 

the contaminant load of runoff and drainage water (McDowell et al., 2014). Furthermore, drainage of wetland 

soils has been shown to reduce water and contaminant storage capacity in the landscape and can increase 

the potential for downstream flooding. The abstraction of surface or groundwater for irrigation disrupts the 

natural water cycle and may stress downstream ecosystems and communities. Irrigation of agricultural lands 

accounts for about 70 percent of ground and surface water withdrawals, and in some regions competition for 

water resources is forcing irrigators to tap unsustainable sources. Irrigation with wastewater may conserve 

fresh water resources but brings the risk of water-borne contaminants in soil and crops (Sato et al., 2013) and 

the accumulation of salts in some environments. 


23

Management 

(global trend) 

Provisioning Regulating Cultural 

Land use change 

(agricultural to urban) 

Decreased biomass, 

decreased availability of 

water for agricultural 

use 

Increased impervious 

surface, decreased 

infiltration, storage, 

soil-mediated water 

regulation 

Decreased natural 

environment 

Increased C 


(increase in change 

of arable to intensive 

grassland) 


(increase in change 

of arable to intensive 

grassland) 

sequestration, greater 

requirement for water, 

stress on ecosystem 

health of downstream 

waterways 

Irrigation (increase) 

Increased biomass over 

dryland agriculture, 

decreased availability of 

water for urban use 

Increased C 

sequestration, but 

decreased filtration 

potential 

Infrastructure alters 

landscape 

Drainage (increase in 

marginal land) 

Decreased soil 

saturation, increased 

biomass, reduction in 

wetlands 

Decreased C 

sequestration, 

denitrification and flood 

attenuation 

Decreased recreational 

potential (e.g. 

ecotourism) 

Table 2.2 Examples of global trends in soil management (Steinfield et al., 2006; Setälä et al., 2014) and their effects on the ecosystem 

services mediated by water. 

The soil management practices to maintain the ecosystem services of food and water security and flow 

regulation within the soil and water cycle are reasonably well established. However, their application is not 

universal and poor management leads to a loss of function. Under climate change scenarios of increased 

climatic variability with more extremes of precipitation, soil functions will be stressed and better soil 

management will be required (Walthall et al., 2012). 

2.4 | Soil as a habitat for organisms and a genetic pool 

Soils represent a physically and chemically complex and heterogeneous habitat supporting a high diversity 

of microbial and faunal taxa. For example, 10 g of soil contains about 1010 bacterial cells of more than 106 

species (Gans et al., 2005), and an estimated 360 000 species of animals are dwellers in soil (Decaëns et al., 

2006). These complex communities of organisms play critical roles in sustaining soil and wider ecosystem 

functioning, thus conferring a multitude of benefits to global cycles and human sustainability. Specifically, 

soil biodiversity is critical to food and fibre production. It is also an important regulator of other vital soil 

services including nutrient cycling, moderation of greenhouse gas emissions, and water purification (Wall et 

al., 2012). It is also recognized that the stocks of soil biodiversity represent an important biological and genetic 

resource for biotechnological exploitation (Brevik and Sauer, 2015). Previous methodological challenges in 

characterizing soil biodiversity are now being overcome through the use of molecular technologies. As a result 

significant progress is currently being made in opening the ‘black box’ of soil biodiversity (Allison and Martiny, 

2008), particularly in assessing the normal operating ranges of soil biodiversity under different soil, climatic 

and land use scenarios. Addressing these knowledge gaps is of fundamental importance, both as an entry 

point to understanding wider soil processes and as a way to gauge the likely consequences of land use or 

climatic change on both biodiversity and soil ecosystem services. 


24

The development of molecular technologies has aided morphological characterisations and allowed 

quantification of stocks and changes in soil biodiversity. This has led to a surge in studies characterizing soil 

biodiversity at different scales – from large landscape-scale surveys to locally focused studies. The largescale 

surveys yield the broader picture, and conclusions are emerging identifying the importance of soil 

parameters in shaping the biodiversity of soil communities (Fierer and Jackson, 2006). In essence, the same 

geological, climatic and biotic parameters that ultimately dictate pedogenesis are also involved in shaping 

the communities of soil biota and thus in regulating the spatial structure of soil communities observed 

over large areas (Griffiths et al., 2011). Locally focused experimentation then typically reveals more specific 

changes in broad taxonomic features with respect, for example, to local changes in land use or climate. 

Many studies have focused on assessing one component of soil diversity, but even greater advances utilizing 

next-generation high throughput sequencing now allow the analysis of ‘whole soil foodwebs’. This permits a 

thorough interrogation of trophic and co-occurrence interaction networks. The challenge is to consolidate 

both approaches at different scales to understand the differing susceptibility of global soil biomes to change. 

Alongside these new developments in assessing biodiversity, it is essential to link the biodiversity 

characteristics measured to specific soil functions. This helps understanding the pivotal roles of soil organisms in 

mediating soil services. The development of stable isotope tracer methodologies (e.g. Radajewski et al., 2000) 

to link substrate utilization to the identified active members in situ serves to clarify the physiological activity of 

these soil organisms. Additionally, improved sequencing techniques are now becoming an increasingly costeffective 

for assessing the biodiversity of functional genes in soils for both eukaryotes and prokaryotes (Fierer 

et al., 2013). This potentially allows a more trait-based approach to understanding soil biodiversity, akin to 

recent approaches applied to larger and more readily functionally understood organisms above-ground. It is 

becoming increasingly apparent that often, as is typical in natural ecosystems, functionality and biodiversity 

co-vary with other environmental parameters. Further manipulative experimentation is required to determine 

the fundamental roles of soil biodiversity versus other co-varying factors in driving soil functionality. 

Clearly, we are learning more and more about how global change affects soil biodiversity and functioning. 

Global-scale syntheses on soil biodiversity are still lacking, but projects such as the Global Soil Biodiversity 

Atlas (European Commission, 2015) are combining information from across the globe and making it publicly 

available. However, much remains to be done. More than 20 years ago, many of these issues were raised (for 

example, in Furusaka, 1993), and to date many of the factors involved have yet to be unravelled. A key barrier 

to achieving syntheses is the lack of concerted soil surveys that address multiple functions using standardized 

methodologies. 

New technologies for soil biodiversity assessment generate large sequence datasets that are typically 

archived in publicly accessible databases. However, morphological datasets remain largely unpublished. 

The best approach to addressing the gaps would be to adopt agreed standard operating procedures for soil 

function measurements (e.g. as developed in the recent EU-funded EcoFINDERS project) and to ensure that 

results are widely accessible. 

Ultimately the new methods are revealing the high sensitivity of changes in soil biological and genetic 

resources to threats such as poor management. We now need to recognize the distinct types of organisms 

found in different soils globally, and to understand their functional roles in order to predict vulnerability of 

these resources to future change. 


25

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3 | Global Soil Resources 

3.1 | The evolution of soil definitions 

The definition of soil has changed over time. Early definitions (Kraut, 1853; Ramann, 1919) emphasized the 

geological or substrate aspect of soil as the upper weathering mantle of the earth’s crust. At the end of the 

19th century, Vasiliy Dokuchaiev formulated the paradigm of soil as a natural body formed by the combined 

effect of five soil-forming factors (climate, organisms, parent material, time and relief). This formulation 

effectively made Dokuchaiev the founder of a new science – pedology. His ideas were translated into English 

and promulgated by Coffey (1912) and Marbut (1921). Jenny (1941) published the equation of soil forming factors 

as independent variables; S = f(cl, o, r, p, t…). Dudal, Nachtergaele and Purnell (2002) added a human factor of 

soil formation, implying that soil is not exclusively a natural body. 

For digital soil mapping, the soil forming factors were modified by McBratney, Mendonça-Santos and 

Minasny (2003) as Sc = f(s, c, o, r, p, a, n…) or Sa = f(s, c, o, r, p, a, n…) where Sc is soil classes, Sa is soil attributes, 

s is the soil or property at a point, and n is the spatial position. Grunwald, Thompson and Boettinger (2011) 

further expanded the factor model to the STEP-AWBH Model by including space and time to infer soil properties 

and their evolution in which the factors of human action, atmosphere, and water are added. 

Defined in the simplest terms, soil is the upper layer of the Earth’s crust transformed by weathering and 

physical/chemical and biological processes. It is composed of mineral particles, organic matter, water, air and 

living organisms organized in genetic soil horizons (ISO, 2013). 

Status of the World’s Soil Resources | Main Report Global Soil Resources 

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3.2 | Soil definitions in different soil classification systems 

The World Reference Base for Soil Resources (FAO, 2014) classifies as soil any material within 2 m of the 

Earth’s surface that is in contact with the atmosphere, but excluding living organisms, areas with continuous 

ice not covered by other material, and water bodies deeper than 2 m. 

In the United States Soil Taxonomy (Soil Survey Staff, 1999) soil is considered to be a natural body comprised 

of solids (minerals and organic matter), liquid, and gases that occurs on the land surface, occupies space, 

and is characterized by one or both of the following: (i) horizons or layers that are distinguishable from the 

initial material as a result of additions, losses, transfers, and transformations of energy and matter; and (ii) the 

ability to support rooted plants in a natural environment. 

In the Russian Classification System (Shishov et al., 2004), soil is defined as a solid-phase natural-historical 

body with a system of inter-related horizons composing a genetic profile and which derives from the 

transformation of the uppermost layer of the lithosphere by the integrity of soil-forming agents. 

French pedologists put emphasis on the spatial aspects of soil as an ‘objet naturel, continu et tridimensionnel’ 

(a natural, continuous and three dimensional object) (AFES, 2008). A related variant considers that “soil in 

nature is a three-dimensional continuum, temporally dynamic and spatially anisotropic, both vertically and 

laterally” (Sposito and Reginato, 1992). 

Urban soils including those ‘sealed’ by concrete or asphalt, strata of composts or other fertile materials 

applied to construct lawns and gardens, superficial layers, mine spoil or garbage heaps are also considered in 

some soil classification systems (Rossiter, 2007). The concept of soils as natural bodies also includes very thin 

films in caves or fine earth patches within desquamation cracks of hard rocks as found in Antarctic endolithic 

soils (Goryachkin et al., 2012) and in underwater soils (Demas, 1993). Thus, the concept of soil becomes very 

broad. Soil scientists have even proposed to extrapolate it to other planets (Targulian et al., 2010). 

3.3 | Soils, landscapes and pedodiversity 

The relationships between soils and landscapes were at the core of the ‘zonality’ concept developed by 

Dokuchaev and tested during his excursion to the Caucasus in 1898. He expressed the concept at the global 

scale in the form of many-coloured soil bands around the Earth. This zonal concept was also used in the United 

States 1938 classification of zonal, azonal, and intrazonal soils (Baldwin, Kellogg and Thorp, 1938). Along with 

zonal ideas, concepts of regularities in local soil patterns emerged. The earliest among these was the concept 

of soil series developed in the United States in 1903 (Simonson, 1952). The work of Neustuev (1931) on soil 

geography further developed the concept of regularities. 

Another set of spatial soil patterns related to topography was recognized by Milne (1935) and Bushnell 

(1945) who proposed the term ‘catena’ (chain) and applied it to soil sequences on the slopes of mountains. 

Different soil catenas in landscapes all over the world were subsequently described and attempts were made 

to inventory them systematically (Sommer and Schlichting, 1997). 

Fridland (1976) gave a new impulse to the theory of the ‘soil/landscape’ relationship by defining the types of 

soil systems related to landforms at different scales (‘soil associations’). The relationships between soils – their 

ingredients, taxonomic distances, geometric shape and kinds of boundaries - were described and for some of 

them mathematic formulas were proposed. 


32

Fridland’s was the first attempt to analyse and quantify the pedological diversity of a territory. The concept 

of soil diversity, or pedodiversity (Ibáñez, Jiménez-Ballesta and García-Álvarez, 1990; Ibáñez et al., 1995; 

McBratney, 1992), opened a new conceptual window in soil science (Ibañez and Bockheim, 2013; Toomanian 

and Esfandiarpoor, 2010). Approaches comprised the description and measurement of either the spatial 

distribution of soils, or their evolutionary stages by indicating rates of soil development. Soil development 

makes a contribution to the spatial heterogeneity of the soil because, together with other agents, soils with 

different evolutionary pathways participate in forming the soil cover and so contribute to the creation of 

specific soilscapes. 

The term ‘pedodiversity’ and many tools for studying pedodiversity were adapted from biology. Pedodiversity, 

for example, can be measured just as biodiversity is measured - by means of special indices showing the 

abundance of species and the taxonomic distances between them. A set of mathematical methods, both 

parametric and non-parametrical, can be applied to quantify soil spatial heterogeneity. 

The pedodiversity concept is an updated, quantification-oriented branch of soil geography. Its advantage 

is its compatibility with GIS and remote sensing technologies and its solid base in mathematics and statistics, 

which leads to a broad applicability in environmental sciences and biology. 

3.4 | Properties of the soil 

Because soils have physical, chemical, mineralogical, and biological characteristics, knowledge of the basic 

sciences of geology, chemistry, physics and biology contributes to understanding basic soil properties. The 

solid inorganic fraction defines the soil’s texture, the amount of sand, silt, and clay. Solid particles are arranged 

into aggregates to form diverse structures by biological, chemical and physical processes. Structure describes 

the size, organization, and shape of the soil aggregates. Consistence and strength are how the soil deforms 

under pressure. Texture and structure influence porosity and bulk density. Gases or solutions occupy the soil 

pores. Soil reaction (pH), redox status, carbon, nutrients, and cation exchange capacity are key chemical 

properties. Secondary clay minerals e.g. smectite, vermiculite, illite, influence the soil physical and chemical 

properties and are the primary source of ionic exchange. The abiotic, inorganic properties create a platform 

for the biotic soil component. 

Properties that are seen or felt are part of the soil morphology. Soil morphology is the object of study both 

in nature and in laboratories – micro morphology – with the help of microscopy and computer tomography. 

Soil colour is influenced by the content and type of organic matter and specific minerals including oxides (e.g. 

Fe oxi-hydroxides), and redox conditions. Horizon and total soil thickness describe internal organization and 

root and moisture availability. 

3.5 | Global soil maps 

Local soil investigations started at the end of the 19th century in Russia (see 3.3. above), but only after World 

War II were efforts geared towards more systematic national soil inventories. The first regional maps were 

produced in the early 1960s for Europe (FAO/UNESCO, 1962) and for Africa (D’Hoore, 1964). 

The development of a global soil map was initiated by the International Soil Science Society in 1960 and 

implemented by FAO and UNESCO between 1971 and 1980, resulting in the FAO-UNESCO Soil Map of the 

World. 1 

1 A digital version of this map is downloadable at: http://www.fao.org/geonetwork/srv/en/resources.get?id=14116&fname=DSMW.zip&access=private 


33

This Soil Map of the World was, from 1995 onwards, systematically updated under the Soil and Terrain 

Database (SOTER) program carried out by FAO, ISRIC and UNEP together with national soil survey services. 

This resulted in several regional updates, including for Latin America and the Caribbean, large parts of Africa, 

and Eastern and Central Europe. In parallel, other organizations, notably the Joint Research Centre (JRC) of the 

European Commision (EC) and the USDA, undertook regional soil updates, while several countries completed 

national soil inventories and maps (China, Brazil, Botswana and Kenya etc.). This updated information was 

harmonized with the digitalized Soil Map of the World and published by a consortium of FAO, IIASA, JRC, 

ISRIC and CAS in 2006 as the Harmonized World Soil Database (HWSD). Although not fully harmonized and 

consistent, the HWSD contains the most up-to-date and comprehensive soil information that is currently 

available. The latest version of this database, giving geo-referenced estimates of twenty soil characteristics, 

is available online. 2 

In 2006, work began on the design and planning for a soil grid of the world at fine resolution (100 m) and 

this became known as GlobalSoilMap. The intent was to integrate the best available data from local and 

national sources and deliver the information online. The format and resolution was to be compatible with 

other fundamental data sets on terrestrial systems (e.g. vegetation, land cover, terrain, remote sensing). The 

initial focus was Africa (Sanchez et al., 2009) and this led to the establishment of the African Soil Information 

System (AfSIS). 3 The technical and logistical complexity of the project has been substantial but good progress 

has been made during the initial research phase of the project and continental coverages are starting to be 

published. 4 A full summary is provided by Arrouays et al. (2014). 

Another, more recent initiative that arose from the GlobalSoilMap effort is Soil Grid 1km 5 which is a 

collection of updatable soil property and class maps of the world at a relatively coarse resolution of 1 km. 

These maps are being produced using state-of-the-art model-based statistical methods: 3D regression with 

splines for continuous soil properties and multinomial logistic regression for soil classes. SoilGrids 1km are 

outputs of a system for automated global soil mapping developed within the Global Soil Information Facilities 

framework. This system is intended to facilitate global soil data initiatives and to serve as a bridge between 

global and local soil mapping (Hengl et al., 2014). 

Information on the availability of global, regional and national soil maps has been summarized by Omuto, 

Nachtergaele and Vargas (2012). The plan for developing the global soil information system was endorsed by 

the Plenary Assembly of the Global Soil Partnership in July 2014 and it is now being implemented. 6 

A simplified global soil map with the major soil groups is given in Figure A 35 (Annex). 

3.6 | Soil qualities essential for the provision of ecosystem services 

Soil functions depend on a number of physical, chemical and biological soil properties that in combination 

determine essential soil qualities. These qualities in turn guarantee that the soil can fulfil its ecological and 

productive services. Soils differ considerably in terms of properties, qualities, limitations and potential. 

Significant changes may occur over very short distances, making environmental and soil monitoring difficult 

(Brammer and Nachtergaele, 2015). 

2 http://www.fao.org/soils-portal/soil-survey/soil-maps-and-databases/harmonized-world-soil-database-v12/it/ 

3 http://www.africasoils.net 

4 http://www.clw.csiro.au/aclep/soilandlandscapegrid/ 

5 http://www.isric.org/content/soilgrids 

6 http://www.fao.org/fileadmin/user_upload/GSP/docs/plenary_assembly_II/pillar4.pdf 


34

Soil management has a considerable effect on how the soil may fulfil its ecosystem services. Mineral 

and organic fertilizer may compensate for poor inherent nutrient conditions in a soil; drainage may remedy 

excessive wetness in soils, or leach salts when these are present; amendments (lime or gypsum) may correct 

very acid or highly sodic soils. However, these interventions always have a cost in terms of labour and inputs, 

and they may also have negative side effects, such as groundwater contamination. 

In this section a number of soil qualities essential for the provision of ecosystem services are discussed and 

related to the major soil groups summarized and illustrated in Annex A 35. 

3.6.1 | Inherent soil fertility 

The capability of a soil to provide sufficient nutrients to crops, grasses and trees is a major quality of soils 

that supports all provisioning services of the ecosystem. Sixteen nutrients are essential for plant growth and 

living organisms in the soil. These fall into two different categories: macronutrients and micronutrients. 

Macronutrients are the most important nutrients for plant development and relatively high quantities 

are required. Macronutrients include: carbon (C), oxygen (O), hydrogen (H), nitrogen (N), phosphorus (P), 

potassium (K), calcium (Ca), magnesium (Mg), and sulphur (S). Micronutrients, on the other hand, are needed 

in smaller amounts, but are still crucial for plant development and growth. Micronutrients include iron 

(Fe), zinc (Zn), manganese (Mn), boron (B), copper (Cu), molybdenum (Mo) and chlorine (Cl). Nearly all plant 

nutrients are taken up in ionic forms from the soil solution as cations or as anions. 

Soil properties directly related to the amount and availability of nutrients in the soil are: (i) soil texture 

(clayey soils contain more nutrients than sandy ones); (ii) the type of clay minerals present (smectitic clays 

absorb more ions than kaolinitic ones); (iii) the soil organic carbon content (more SOC corresponds with a 

larger amount of nutrients); and (iv) the cation exchange capacity that corresponds to the total of Ca, Mg, 

K, Na (basic ions) and Al and H (acidic ions) exchangeable with the soil solution. A large amount of available 

nutrients is present in Vertisols, Chernozems (Borolls), Kastanozems (Ustolls) and Phaeozems (Udolls). Also 

volcanic soils (Andosols) and alluvial soils (Fluvisols/Fluvents) generally have a large nutrient content. On 

the other hand, sandy soils (Arenosols/Psamments) and highly leached soils (Ferralsols/Oxisols and Acrisols/ 

Ultisols) generally have a small nutrient content. 

The amount of nutrients that a soil can provide to plants within the growing season represents a limit to 

nutrient mining. Nutrient mining occurs when crops take out a high proportion of the nutrients available in 

the soil, leaving a nutrient imbalance that threatens the sustained provision of food and ecosystem services. 

These challenges are discussed in Section 6.8. Figure 3.1 illustrates an estimation of the nutrient availability in 

soils globally based on information contained in HWSD. 

Soil depth to a hard or an impermeable layer is a vital factor that determines the capability of roots to take 

hold and determines the total volume of nutrients and water available to crops and vegetation. Soils tend to 

be deeper when strong weathering conditions prevail over a long period and wherever the parent material is 

readily weathered. Typical soils include Ferralsols and Nitisols). Shallow soils often occur in mountainous areas 

(Leptosols) and in dry areas characterized by indurated layers of silica, calcium carbonate or gypsum (Durisols/ 

Durids, Calcisols/Calcids and Gypsisols/Gypsids). Each plant type has its own ideal rooting conditions. Tubers 

are the most sensitive to soil depth and volume limitations (Fischer et al., 2008; Grossnickle, 2005; Unger 

and Kaspar, 1994; McSweeney and Jansen, 1984; Myers et al., 2007). Figure 3.2 illustrates global soil rooting 

conditions 


35

No or slight constraints Moderate constrains Severe constraints Very severe constraints Mainly non-soil Permafrost area Water bodies 

Figure 3.1 Nutrient availability in soils. Source: Fischer et al., 2008. 

The soil pH is a measure of its hydrogen ion concentration and indicates the acidity or alkalinity of the 

soil. Optimum availability of nutrients occurs around pH=6.5. Toxic concentrations of H and Al occur when 

the pH drops below 5.5. Values of pH above 7.2 indicate an alkaline reaction and may be symptomatic for the 

immobilization of nutrients. Very high pH values over 8.5 result in the dispersion of the soil particles and a 

collapse of structure. High rainfall results in more acid soils (Ferralsols/Oxisols, Alisols, Plinthisols, Acrisols/ 

Ultisols, Podzols/Spodosols), while drier conditions often lead to the accumulation of Gypsum (Gypsisols/ 

Gyspsids) or other less soluble salts (Silicon and Calcium Carbonate) in Durisols/Durids and Calcisols/Calcids. 

The soil pH is also important to the characterization of soil threats to ecosystem services such as acidification 

(section 6.4) and sodification (Section 6.5). A global map of soil pH is given in Section 6.4. 

No or slight constraints Moderate constraints Severe constraints Very severe constraints Mainly non-soils Permafrost area Water bodies 

Figure 3.2 Global soil rooting conditions. Source: Fischer et al., 2008. 


36

Accumulation of water soluble salts: soils in relatively dry areas are often characterized by the accumulation 

of water-soluble salts (NaCl, Na 2 

SO 4 

, etc.) and of less water-soluble salts (CaCO 3 

, CaSO 4 

). These salts may 

form indurated layers that limit the soil depth available for roots. This accumulation is often a natural process 

resulting in soils such as Solonchaks/Salids, Calcisols/Calcids, Gypsisols/Gypsids, Durisols/Durids. In irrigation 

schemes, which are most commonly developed in dry areas, the problem may be human-induced and made 

worse by the use of saline irrigation water, by insufficient leaching/drainage, or by the conversion to irrigation 

of soils formed from marine sediments. Most salt-affected soils have moderate to severe limitations for crop 

production. Section 6.5 deals specifically with salinization and sodification problems. 

Toxic elements and other soil fertility problems: some toxic elements such as aluminium occur naturally 

in acid soils. The parent material may also be a natural source of undesirable elements (for instance cadmium) 

that may be a problem for human and animal health. Some soils have a high phosphorus adsorption ratio 

(Andosols and Nitisols/Kandi subgroups) that make P fertilization cumbersome. Atmospheric deposition of 

toxic elements may also contaminate soils as discussed in section 4.4. 

3.6.2 | Soil moisture qualities and limitations 

The moisture stored in or flowing through the soil affects soil formation, its structure and stability, and 

erosion run-off. Soil moisture is of primary concern with respect to plant growth. The depth of the groundwater 

table and the availability of oxygen in the soil also affect soil ecosystem functions. The physical properties of 

soils (texture, structure, porosity, drainage class, permeability) are of prime importance in this respect. 

The capacity to store water and moisture in a soil is largely determined by its texture, structure, 

organic carbon content and depth. Soil moisture provides a buffer for crops during dry periods and is a builtin 

safeguard against run-off and erosion. Ecological functions of this parameter are discussed in Chapter 7. 

High soil moisture capacities are typical for deep clayey soils, rich in organic matter and containing modest 

amounts of CaCO 3 

(Chernozems, Cambisols). The lowest soil moisture capacities are encountered in sandy 

soils (Arenosols) or very shallow soils (Leptosols). Very high soil moisture storage occurs in volcanic soils 

(Andosols) and in many peat soils (Histosols). Figure 3.3 illustrates the distribution of different soil moisture 

storage classes globally. 

Oxygen availability is a critical factor for plant growth. Inadequate oxygen supply to the roots leads to the 

formation of an underdeveloped root system which is not able to provide sufficient nutrients and water to 

the plant. Oxygen availability is basically defined by drainage characteristics of soils related to soil type, soil 

texture, soil phases and terrain slope, all of which play an important role in determining the proportion of gases 

and water into the soil. Soil phases define specific soil and terrain characteristics. Gleysols/Aquic suborders, 

Stagnosols and Plinthosols often suffer from temporary saturation with groundwater or rain water, resulting 

in poor oxygen availability for part of the year. Oxygen availability can be improved by farming practices (e.g. 

adapted tillage) and by farming inputs such as artificial drainage (Crawford, 1992; Erikson, 1982; Fischer et al., 

2008). 

3.6.3 | Soils properties and climate change 

Soils are both affected by and contribute to climate change. The carbon that is fixed by plants is transferred 

to the soil via dead plant matter including dead roots and leaves. This dead organic matter creates a substrate 

which soil micro-organisms respire back to the atmosphere as carbon dioxide or methane depending on 

the availability of oxygen in the soil. Some of the carbon compounds are easily digested and respired by the 

microbes, resulting in a relatively short residence time. Others become chemically and/or physically stabilised 

in soils and have longer residence times (as described in Chapter 2). Soil organic carbon can also be thermally 

decomposed during fire events and returned to the atmosphere as carbon dioxide. Remaining charred material 

can persist in soils for long periods (Lehmann et al., 2015). 


37

Average water capacity class (mm/m) 

> 150 100 - 150 50 - 100 < 50 

Figure 3.3 Soil Moisture storage capacity. Source: Van Engelen, 2012. 

Soil organic carbon improves the physical and chemical properties of the soil by increasing the cation 

exchange capacity and the water-holding capacity. It also contributes to the structural stability of soils by 

helping to bind particles into aggregates. Soil organic matter (SOM), of which carbon is a major part, holds 

a great proportion of nutrients, including trace elements, which are of importance to plant growth. SOM 

mitigates nutrient leaching and contributes to soil pH-buffering capacity. It is widely accepted that the 

organic matter content of the soil is a major factor contributing to soil functions, including that of organic C 

storage, which has important feedbacks with the Earth’s climate system (Chapter 2). 

A large organic carbon content is found in peat soils (Histosols/Histisols), in volcanic soils (Andosols/ 

Andisols) and in steppe soils (Chernozem/Borolls, Kastanozems/Ustolls and Phaeozems/Udolls). Large 

organic carbon contents are not always indicative of fertile soils because carbon may also accumulate under 

wet and cold conditions as in Podzols/Spodosols and Cryosols/Gelisols, and in some hydromorphic soils such 

as Gleysols. Changes in SOC represent one of the major soil threats – see the discussion in section 6.2. The 

global distribution of soil organic carbon is given in Figure 3.4. 

Crysols/Gelisols are soils which are frozen for a large part of the year. In taiga areas they often occur 

together with Histosols. Global warming in these areas will have a significant effect by allowing agriculture 

to move more northwards. However, mineralization of organic carbon may be accelerated, with negative 

consequences for GHG release. 

3.6.4 | Soil erodibility and water erosion 

The susceptibility of a soil to water erosion is primarily determined by the erosive potential of the rainfall, 

the slope of the land surface and position of the soil in the catchment, and the vegetative cover on the soil 

surface. Soil erodibility refers to the susceptibility of soil to erosion by water and is an important secondary 

control on the intensity of water erosion. Most clay-rich soils (e.g. Vertisols with the exception of erodible 

self-mulching forms) have a high resilience because they are resistant to detachment. Coarse textured, sandy 

soils (e.g. Arenosols/Psamments) are also resilient because of low runoff even though these soils are easily 

detached. Medium textured soils, such as silt loam soils are only moderately resistant to erosion because 

they are moderately susceptible to detachment and they produce moderate runoff. Soils having a high silt 

content are the most erodible of all soils. They are easily detached, tend to crust and produce high rates of 

runoff. Organic matter reduces erodibility because it reduces the susceptibility of the soil to detachment, 


38

and increases infiltration, which reduces runoff and thus erosion. Soil structure affects both susceptibility to 

detachment and infiltration. Permeability of the soil profile affects erodibility because it affects runoff. Past 

management or misuse of a soil (e.g. by intensive cropping) can increase a soil’s erodibility, for example if the 

subsoil is exposed or if the organic matter has been depleted, or where the soil’s structure has been destroyed 

or soil compaction has reduced permeability. Section 6.1 discusses soil erosion by water in more detail. Soil 

erodibility worldwide, as characterized by the k factor in the RUSLE equation, is represented in Figure 3.5. 

Prepared by R. Hiederer 

Figure 3.4 Soil Organic Carbon pool (tonnes C ha -1 ). 

3.6.5 | Soil workability 

Soil workability refers to the ease of tillage, which depends on the soil’s interrelated characteristics of texture, 

structure, organic matter content, etc., on the soil’s gravel content, and on the presence of continuous hard 

rock at shallow depth. Depending on the soil characteristics, soil workability also varies with the soil moisture 

content. Some soils are easy to work regardless of the moisture content, but other soils – such as Vertisols 

- can be worked only at a specific moisture status. This is true especially for farming systems employing 

manual cultivation methods or using only light machinery. Soil workability is also related to the type of soil 

management adopted. While low and intermediate input farming systems mainly face constraints related to 

soil texture and soil structure, high-level input mechanized farming systems mainly face constraints related to 

irregular soil depth and stony and rocky soil conditions. Indeed, the use of heavy field equipment is not possible 

on stony soils or on soils characterized by irregular soil depth. This factor can prevent soil degradation, for 

example by compaction (Earl, 1997; Fischer et al., 2008; Müller et al., 2011; Rounsevell, 1993). Figure 3.6 shows 

the distribution of the constraints to soil management and food production due to soil workability worldwide. 


39

Undefined < 0.06 

0.06 - 0.12 0.12 - 0.18 0.18 - 0.24 

0.24 - 0.3 0.3 - 0.36 0.36 - 0.42 0.42 - 0.48 0.48 - 0.54 

< 

0.54 

Figure 3.5 Soil erodibility as characterized by the k factor. Source: Nachtergaele and Petri, 2011. 

3.6.6 | Soils and ecosystem goods and services 

Figure 3.7 illustrates the suitability of soils for supporting crops. The evaluation is based on soil health but 

excludes climatic considerations (except for low temperatures). In Table 3.1, the contribution of the main soil 

types to major ecosystem services (food security, climate regulation, water regulation and socio-cultural 

provisions) is estimated at a scale from zero to five. The ratings are based on soil characteristics and quality as 

measured by: suitability for growing crops; organic carbon content; water holding capacity; and capacity to 

support infrastructure and store archaeological remains. 

No or slight constraints Moderate constraints Severe constraints Very severe constraints Mainly non-soils Permafrost areas Water bodies 

Figure 3.6 Soil workability derived from HWSD. Source: Fischer et al., 2008. 


40

Figure 3.7 Soil suitability for cropping at low input, based on the global agro-ecological zones study. 

Source: Fischer et al., 2008. 


41

Ecosystem Services 

Reference 

Soil Groups 

Food Climate Water Cultural SUM Major Service 

Histosols 2 5 5 3 15 Climate Change 

Anthrosols 5 5 5 4 19 Food Security 

Technosols 1 3 2 4 10 Infrastructure 

Cryosols 0 5 2 3 10 Climate Change 

Leptosols 1 1 2 1 5 Water runoff 

Vertisols 4 2 3 1 10 Food Security 

Solonetz 1 1 1 1 4 Very few 

Solonchaks 1 1 1 1 4 Very few 

Podzols 1 3 1 1 6 Biomass 

Ferralsols 2 4 3 1 10 Biomass 

Nitisols 4 3 4 1 12 Food Security 

Plinthosols 2 1 2 1 6 Biomass 

Planosols 1 1 1 1 4 Very few 

Gleysols 2 1 3 1 7 Food Security 

Stagnosols 2 1 3 1 7 Water storage 

Andosols 4 3 5 1 13 Food Security 

Chernozems 5 4 4 1 14 Food Security 

Kastanozems 3 4 2 1 10 Food Security 

Phaeozems 4 4 3 1 12 Food Security 

Umbrisols 3 3 3 1 10 Water runoff 

Durisols 1 1 1 1 4 Very few 

Calcisols 1 1 2 1 5 Very few 

Gypsisols 1 1 1 1 4 Very few 

Retisols 2 1 2 1 6 Biomass 

Acrisols 2 1 2 1 6 Food Security 

Lixisols 2 1 2 1 6 Food Security 

Alisols 1 1 2 1 5 Biomass 

Luvisols 3 2 2 1 8 Food Security 

Cambisols 3 2 3 1 9 Food Security 

Regosols 2 1 1 1 5 Biomass 

Arenosols 1 1 1 1 4 Biomass 

Fluvisols 4 2 4 2 12 Food security 

Wassents 0 2 2 1 5 Very few 

Prepared by R. Hiederer, JRC 

Table 3.1: Generalized ecosystem service rating of specific soil groups (WRB) 7 

7 Soil Taxonomy equivalents given in the Annex, except for Wassents that are a suborder in Soil Taxonomy 


42

3.7 | Global assessments of soil change - a history 

Global assessments of soil and land degradation started more than 40 years ago, but have until now not 

achieved a clear answer on where soil degradation takes place, what impact it has on the population, and 

what the cost to governments and land users would be if the decline in soil, water and vegetation resources 

continued unabated. Although institutional, socio-economic and biophysical causes of soil degradation have 

been identified locally in many case studies, these have seldom been inventoried systematically at national 

or regional level. Much of the investment in land reclamation and rehabilitation during recent years has been 

driven by donor interest to fund action, rather than research to understand the scope of the problem. Even 

knowledge about what works and what does not work in combating soil degradation is scanty, and there has 

been little systematic investigation. In recent years, however, the World Overview of Conservation Approaches 

and Technologies (WOCAT) consortium has begun to make a substantial contribution through its systematic 

collection of information on sustainable soil and water conservation practices and their impacts. 

The first comprehensive assessment of global soil degradation was based on expert opinion only. This 

was the GLobal Assessment of human-induced SOil Degradation - GLASOD, published by UNEP/ISRIC (Oldeman, 

Hakkeling and Sombroek, 1991). The Land Degradation Assessment in Drylands project (LADA) was launched by 

GEF, implemented by UNEP and executed by FAO between 2006 and 2011 in support of the UNCCD. LADA 

developed an approach based on remotely-sensed NDVI data (the Global Land Degradation Assessment – GLADA). 

The project also used an ecosystems approach that brought together and interpreted information from preexisting 

and newly developed global databases to inform decision makers on all aspects of land degradation 

at a global scale (GLADIS: the Global LAnd Degradation Information System). 

During this period other important and broader environmental assessments took place, notably the 

Millennium Ecosystem Assessment (MA, 2005) and the periodical review of the State of the Environment by UNEP 

with the GEO- reports (UNEP, 2012). FAO published a State of Land and Water (SOLAW) in 2011. The Economics of 

Land Degradation (ELD) initiative (ELD, 2015) provided in 2015 a first estimate of the cost of land degradation 

at global scale based on rather scattered and uncertain information. The annual economic losses due to 

deforestation and land degradation were estimated at EUR 1.5–3.4 trillion in 2008, equaling 3.3–7.5 percent of 

the global GDP in 2008. All of these studies used the results of one of the three global inventories: GLASOD, 

GLADA or GLADIS which are discussed in more detail below. 

3.7.1 | GLASOD: expert opinion 

An expert consultation on soil degradation convened by FAO and UNEP in Rome in 1974 recommended 

that a global assessment be made of actual and potential soil degradation. This assessment, which was 

conducted in collaboration with UNESCO, WMO and ISSS, was based on the compilation of existing data 

and the interpretation of environmental factors influencing the extent and intensity of soil degradation. 

The assessment considered such environmental factors as climate, vegetation, soil characteristics, soil 

management, topography and type of land utilization. The results of this assessment were compiled as a 

world map of soil degradation. During the next four years FAO, UNESCO and UNEP developed a provisional 

methodology for soil degradation assessment and prepared a first approximation study identifying areas 

of potential degradation hazard for soil erosion by wind and water, salinization and sodification. Maps 

at a scale of 1:5 M covering Africa north of the equator and the Middle East were prepared (FAO/UNEP/ 

UNESCO, 1979). These first efforts were then scaled up into the Global Assessment of Human Induced Soil 

Degradation Project or GLASOD. The project was initiated by UNEP. It had a duration of 28 months and 

was executed by ISRIC. In order to cover the whole world, 21 regions and individual countries were defined 

and experts on these regions were asked to prepare detailed maps of soil degradation. More than 250 soil 

scientists and environmentalists cooperated in this project (Oldeman, Hakkeling and Sombroek, 1991). 

The global results of the GLASOD project are available online. 8 

8 http://www.isric.org/UK/About+ISRIC/Projects/Track+Record/GLASOD.htm 


43

A regional follow-up in Southeast Asia resulted in a more detailed database for that region: ASSOD (Van 

Lynden and Oldeman, 1997). 

Since its publication, some expert opinion has faulted GLASOD, questioning the objectivity and 

reproducibility of an assessment based on expert opinion as an assessment approach (Sonneveld and 

Dent, 2007). However, at the time GLASOD was developed there were few alternatives available, especially 

given the overall lack of remotely sensed data at the time. Even today the criticism seems unwarranted as 

remotely sensed techniques and most modelling approaches have so far failed to come up with more useful 

assessments. GLASOD results are presented in Figure 3.8. 

Figure 3.8 GLASOD results. Source: Oldeman, Hakkeling and Sombroek, 1991. 


44

3.7.2 | LADA-GLADIS: the ecosystem approach 

The first approaches of LADA (see 3.7 above) used remotely-sensed NDVI data to prepare the Global Land 

Degradation Assessment – GLADA (Bai et al., 2008). However, this was soon superseded by a complementary 

approach that focused on the actual status and trends of land resources in terms of six factors: biomass, 

water resources, soil health 9 , above-ground biodiversity, and economic and social provisions that contribute 

to ecosystem goods and services (Figure 3.9). The evaluation was based on interpretation of global databases 

available in the public domain, using documented algorithms to achieve a rating for each of the six factors 

in terms of status and trends. In order to map the various aspects, a special ‘global land use system’ was 

developed (Nachtergaele and Petri, 2011) which allowed cause and effect to be linked. Results were presented 

in radar diagrams (Figure 3.10) that showed the variability of ecosystem services provided as a function of land 

use and the need for trade-offs between different factors related to ecosystem goods and services. The GLADIS 

system is accessible on-line at: 

http://www.fao.org/nr/lada/index.php?option=com_content&view=article&id=161&Itemid=113&lang=en 

An example of an output for global soil compaction is shown in Figure 3.10. 

Criticism of the GLADIS system focused on the unreliability of some of the global databases used and on 

questions about the downscaling relationships that were developed at local scale (such as the RUSLE). For 

the specific factor - soil health - the absence of an assessment of wind erosion is certainly a limitation, while 

the fact that no difference is made between ‘natural’ and ‘human induced’ soil erosion is also confusing. These 

weaknesses have been recognized and should be corrected where possible during the further development of 

the GLADIS information system which is pending. 

Forest 

Urban 

Biomass (Acc) 

Biomass (Acc) 

Soc./Cult. Benefit 

Biomass (Ann) 


Biomass (Ann) 

Economic benefit 




Biodiv. 

Water Q/Q 

Biodiv. 

Water Q/Q 

Agriculture 

Biomass (Acc) 


Biomass (Ann) 



Biodiv. 

Water Q/Q 

Figure 3.9 Example of the effect Figure of land 3.9: Example use on of indicative the effect of land factors use on for indicative ecosystem factors for goods ecosystem and goods services and services 

9 The soil health status was obtained by comparing the soil suitability for the actual land use. The soil health trend was based on a combination of ratings for the 

risk of erosion by water, the soil compaction risk, a nutrient balance, and the soil contamination and soil salinization risks. 


45

Ocean / Seas (0) 

Inland water 

Figure 3.10 Soil compaction risk derived from intensity of tractor use in crop land and from livestock density in grasslands. Source: 

Nachtergaele et al., 2011 

3.7.3 | Status of the World’s Soil Resources 

The present book – The Status of the World’s Soil Resources - takes a different approach from the ones 

described above by focusing on well documented and peer reviewed research data on soil degradation 

processes, status and trends in scientific literature at all levels. It also draws attention to the uncertainty of 

estimates made. 

The quantity and quality of information on soil degradation is shown to be very variable in different regions. 

Some regional statements - Africa, Eurasia, Near East, Latin America - still rely on GLASOD or ASSOD. For other 

regions, such as North America, no regional harmonized approach has been undertaken. Only the EU and the 

South West Pacific have made progress in establishing new regional updated approaches. 

The report also shows the great differences that exist in data and data availability on soil resources and 

soil change information at national level. Systematic sampling/surveying and monitoring does take place for 

selected major land uses (forests, arable lands) in most EU countries, the United States and Canada, China, 

Australia and New Zealand. However, results are not always made available in the public domain. The progress 

in digital soil mapping may help more countries to produce harmonized data and to make the information 

public. 

The data presented in this book constitute a baseline inventorying the documented knowledge at a point 

in time: 2015. Future progress can thus be measured against this baseline. 


46

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49

4 | Soils and Humans 

4.1 | Current land cover and land use 

The new Global Land Cover Share database (Latham et al., 2014) includes eleven global land cover layers, 

each representing the major land cover classes defined by the FAO and SEEA legend (Weber, 2010). 

Analysis of the database indicates that of the global land mass, artificial surfaces occupy 0.6 percent, 

croplands 12.6 percent, grasslands 13.0 percent, tree-covered areas 27.7 percent, shrub-covered areas 9.5 

percent, herbaceous vegetation 1.3 percent, mangroves 0.1 percent, sparse vegetation 7.7 percent, bare soils 

15.2 percent, snow and glaciers 9.7 percent and inland water bodies 2.6 percent. 

The intensity of each land-cover type varies substantially across the globe according to numerous factors, 

including soils, altitude, climatic conditions and anthropogenic influences. For example, while cultivated land 

is less than 10 percent in most African regions, it accounts for more than 25 percent of the land in the Asia 

region. A land cover map is given in Figure 4.1. Summary statistics by region, derived from the respective GIS 

layers are given in Figure 4.2. In the following discussion, attention is focused on three main land cover classes: 

cropland, grasslands/grazing lands and forests. The management of these three classes has large impacts on 

soils and ecosystem services. The presence of artificial surfaces is treated in more detail in Section 6.7. More 

than 25 percent of the land mass carries almost no vegetation because of climatic factors (glaciers, deserts) or 

topographic or soil conditions. 

Status of the World’s Soil Resources | Main Report Soils and Humans 

50

75% crops Mixed >50% artificial 50-75% grass/shrub 50-75% forest 50-75% crops > 50% non vegetated >75% grass/shrub >75% forest 

Water 

Figure 4.1 Global Land Cover. Source: Latham et al., 2014. 

Figure 4.2 Distribution of land cover in different regions. Source: Latham et al., 2014. 


51

Cropland 

SOLAW (FAO, 2011) established that the cultivated land area in terms of per capita use in 2000 was highest 

in Australia (more than 2.2 ha per person), followed by North America and Eastern Europe and Russia (about 

0.7 ha per person). In contrast, current cultivated land used per capita is only 0.2 ha in Western Europe and in 

most less developed countries. 

By dividing the current cultivated land by the projected populations, the anticipated cultivated land area 

per capita in 2050 can be estimated. In the more developed countries, the cultivated area per capita would 

change little. In less developed countries, the cultivated area per capita is expected to halve to 0.1 ha by 2050, 

unless there is further expansion of the cultivated area. 

Further characterization of cropland and land use at a global scale by remote sensing is difficult because: 

1. The spatial extent of croplands is highly variable between and within nations. Cropland characteristics 

such as field size can be highly variable, even for the same crop type. Spatial extent of cropland depends 

on a host of factors, including the historical, political, social and technological context of agricultural 

development as well as natural factors such as landscape patterns. 

2. Patterns of agricultural intensification – for example, the use of fertilizer – vary greatly, especially 

between developed and developing nations. 

3. Each crop type has a specific growth phenology and structure, with significant seasonal variation 

between and even within individual crop types. 

4. Cropland can be confused with natural vegetation cover types – for example, surveys may confuse 

cereal grains with tall-grass prairie (Pittman et al., 2010). Better cropland information – in terms of 

both its extent and the purpose and intensity of its use – is vital to understanding soil change and 

to formulating adequate responses. Special attention should be paid to irrigated agriculture in 

developing countries, which covers about one-fifth of all arable land, and accounts for 47 percent of all 

crop production and almost 60 percent of cereal production (Nachtergaele et al., 2011). 

Grazing lands 

Grazing lands, including sown pasture and rangeland with various coverage (grasslands, bush/shrublands), 

are among the largest ecosystems in the world and contribute to the livelihoods of more than 800 million 

people. They are a source of goods and services such as food and forage, energy and wildlife habitat, and also 

provide carbon and water storage and watershed protection for many major river systems. Grasslands are 

also important for in situ conservation of genetic resources. Of a total of 10 000 species, only 100 to 150 

forage species have been cultivated, but many more hold potential for sustainable agriculture. Estimates of 

the proportion of the Earth's land area covered by grasslands vary between 20 and 40 percent, depending on 

the definition. Those differences are due to a lack of harmonization in the definition of grasslands. 

There has been a significant reduction of pasture in Eastern Africa, partially because large grassland areas 

have been destroyed or converted to agricultural land. In South America, pastures have been lost because 

of conversion to soybean cultivation. In Europe there has been a gain in grazing lands because European 

policies such as the ‘set-aside’ measures oblige farmers to leave a portion of their agricultural land in fallow as 

a condition for benefiting from direct payments (Suttie, Reynolds and Batello, 2005). 

Forests 

In 2010, forests covered about 28 percent of the world’s total land area. Deforestation affected an estimated 

13 million ha per year between 2000 and 2010. Net forest loss was, however, considerably less – 5.2 million 

ha per year – as losses were compensated by afforestation and some natural expansion (FAO, 2014a). Most 

deforestation takes place in tropical countries, whereas most developed countries with temperate and 

boreal forest ecosystems – and more recently, countries in the Near East and Asia – are experiencing stable 


52

or increasing forest areas. Between 1990 and 2010, the amount of forest land designated primarily for the 

conservation of biological diversity increased by 35 percent, indicating a political commitment to conserve 

forests. These forests now account for 12 percent of the world’s forests. 

Approximately 13.2 million people worldwide are formally employed in the forestry sector. Many more 

depend directly on forests and forest products for their living. In developing countries, wood-based fuels are 

the dominant source of energy for more than two billion mostly poor people. In Africa, over 90 percent of 

harvested wood is used for energy. Wood accounts for 27 percent of total primary energy supply in Africa, 

13 percent in Latin America and the Caribbean and five percent in Asia and Oceania. However, it is also 

increasingly used in developed countries with the aim of reducing dependence on fossil fuels. For example, 

about 90 million people in Europe and North America now use wood energy as the main source of domestic 

heating (FAO, 2014a). 

Conclusion 

Land cover and land use are essential factors to understand soil change. In particular, better cropland 

information, in terms of extent, purpose and intensity of use, is vital to understanding soil change and to 

formulating adequate responses. 

4.2 | Historical land cover and land use change 

Since the early days of agriculture, human activity has altered vegetation cover and soil properties. ‘Land 

use change’ or ‘land cover change’ typically refers to changing from one type of vegetation cover to another 

(e.g forest to pasture, natural grassland to cropland). Although the terms land use change and land cover 

change are often used interchangeably, ‘land use’ is more typically used to refer to management within a land 

cover type. Land use is thus “characterised by the arrangements, activities and inputs people undertake in 

a certain land cover type to produce, change or maintain it" (FAO/UNEP, 1999). Land use change has been 

accelerated by migration and population increase as food, shelter, and materials are sought and acquired. It 

is estimated that humans have directly modified at least 70 million km 2 , or >50 percent of Earth’s ice-free land 

area (Hooke, Martín-Duque and Pedraza, 2012). 

For a long period of human activity, until about a thousand years ago, cropland and pasture occupied less 

than one to two percent each of the global ice-free land area (based on a range of data sources in Klein Goldewijk 

et al., 2011 and depicted in Figure 4.3; also see Ramankutty, Foley and Olejniczak, 2002). Subsequently, as the 

population centres of Europe, South Central Asia and Eastern Asia expanded, more land was converted from 

natural vegegation to cultivated lands. Cover of croplands and pastures was about two to four percent each 

by 1700 (Klein Goldewijk et al., 2011). By 1900, agriculture had further expanded in these areas, and spread to 

North America. Since 1900 rapid expansion has continued, including the arable areas of South America, Africa 

and Australia. As a result, today nearly all soils and climates suitable for cultivation in industralized countries 

are in use for crop production. In some of these countries, cropland expansion has been reversed in recent 

years, as with the EU set-aside programme. South America and Africa continue to convert land use to crop 

production. By 2000, global cropland cover had reached 11 percent and pasture cover 24 percent, according to 

Klein Goldewijk et al. (2011) based on FAO statistics. 


53

Prepared by P. Reich 

Figure 4.3 Historical land use change 1000 – 2005. Source: Klein Goldewijk et al., 2011. 

The net loss of natural land has been dominated by loss of tropical forests (3.3 million km 2 ), tropical 

grasslands (6.8 million km 2 ) and temperate grasslands (5.5 million km 2 ). Quantification from satellite imagery 

of global forest change over the period 2000-2012 shows that tropical deforestation remains the predominant 

source of losses (Hansen et al., 2013). However, there has been a reduced rate of deforestation in some regions 

over the last decade, most notably in Brazil. This is coupled with a rising rate of afforestation in some areas in 

recent decades, notably in Europe and the United States, and more recently in China, Vietnam and India (FAO, 

2013). 

4.3 | Interactions between soils, land use and management 

Many soils are subject to some degree of direct or indirect human disturbance. However, distinguishing 

natural from direct and indirect human influence is not always straightforward (Smith, 2005). Nonetheless, 

some human activities have clear direct impacts. These include land use change, land management, land 

degradation, soil sealing, and mining. The intensity of land use also has a great impact on soils. Soils are also 

subject to indirect impacts arising from human activity, such as acid deposition (for example, sulphur and 

nitrogen) and heavy metal pollution. In this section, we report the state-of-the-art understanding and the 

knowledge gaps concerning these impacts on soils. 

4.3.1 | Land use change and soil degradation 

Land cover change (Section 4.2), for example from forest or natural grassland to pasture or cropland, 

removes biomass and disturbs soils. This in turn leads to loss of soil carbon and other nutrients and to changes 

in soil properties and in soil biodiversity. Some land cover conversions – for example, afforestation after 

abandonment of cropland – can result in increases of soil carbon and nutrients. Land use that does not result 

in a change of cover, such as forest harvest and regrowth, or increasing grazing intensity, can nonetheless 

result in degradation of soil properties. 


54

Degrading land covers approximately 24 percent of the global land area (35 million km 2 ). 23 percent of 

degrading land is broadleaved forest, 19 percent needle-leaved forests, 20–25 percent rangeland (Bai et al., 

2008). The scale and nature of the changes are highly variable with type of land cover change, climate, and 

method of vegetation removal (e.g. land clearing fires, mechanical harvest). This section focuses on metaanalyses 

of field data and global model results. The effects of land use changes within agricultural lands are 

dealt with in Section 4.3.2. 

Impacts of land cover change 

Wei et al. (2014) collated observations from 119 publications of 453 paired or chrono-sequential sites in 36 

countries where tropical, temperate, and boreal forests were converted to agricultural land. The SOC stocks 

were corrected for changes in soil bulk density after land-use change and only SOC in the upper 0–30 cm was 

considered. The SOC stocks decreased at 98 percent of the sites by an average of 52 percent in temperate 

regions 41 percent in tropical regions and 31 percent in boreal regions. The decrease in SOC stocks and the 

turnover rate constants both varied significantly according to forest type, cultivation stage, climate and 

soil factors. A meta-analysis (Guo and Gifford, 2002) of 74 publications across tropical and temperate zones 

showed a decline in soil C stocks after conversion from pasture to plantation (−10 percent), native forest to 

plantation (−13 percent), native forest to crop (−42 percent), and pasture to crop (−59 percent). Soil C stocks 

increased after conversions from native forest to pasture (+8 percent), crop to pasture (+19 percent), crop to 

plantation (+18 percent), and crop to secondary forest (+53 percent). Broadleaf tree plantations placed onto 

prior native forest or pastures did not affect soil C stocks whereas pine plantations reduced soil C stocks by – 12 

to – 15 percent. In this study, soil depth varied from less than 30 cm to more than 100 cm and was not adjusted 

to account for changes in bulk density with land use change. 

In a meta-analysis of 385 studies on land use changes in the tropics (Don, Schumacher and Freibauer, 2011), 

SOC decreased when primary forest was converted to cropland (-25 percent), perennial crops (-30 percent) 

and grassland ( -1 2 percent). SOC increased when cropland was afforested (+29 percent) or under cropland 

fallow (+32 percent) or converted to grassland (+26 percent). Secondary forests stored 9 percent less SOC than 

primary forests. Relative changes were equally high in the subsoil as in the surface soil (Don, Schumacher and 

Freibauer, 2011). In this study, SOC stocks were corrected to an equivalent soil mass and sampling depth was 

on average 32 cm. 

The response of soil organic carbon (SOC) to afforestation in deep soil layers is still poorly understood. Shi 

et al. (2013) compiled information on changes in deep SOC (defined as at least 10 cm deeper than the 0–10 

cm layer) after afforestation of croplands and grasslands (total 63 sites from 56 literature). The responses of 

SOC to afforestation were slightly negative for grassland, and significantly positive for cropland. The SOC in 

soil depth layers (up to 80 cm) was reduced after afforestation of grassland but not significantly. By contrast, 

conversion of cropland to forests (trees or shrubs) increased SOC significantly for each soil layer up to 60 cm 

depth. 

Poeplau et al. (2011) compiled 95 studies conducted on conversion in temperate climates. One finding 

was that topsoil (0-30 cm) SOC decreases quickly (~20 years) when cropland is established on grassland (-32 

percent) or forest (-36 percent). By contrast, long lasting (> 120 years) sinks are created through conversion 

of cropland to forest (+16 percent) or grassland (+28 percent). Afforestation of grassland did not result in 

significant long term SOC stock trends in mineral soils, but did cause a net carbon accumulation in the labile 

forest floor (e.g. 38 Mg ha -1 over 100 years). However, this carbon accumulation cannot be considered as an 

intermediate or long-term C storage since it may be lost easily after disruptions such as fire, windthrow or 

clear cut (Poeplau et al., 2011). 

Peatlands (organic soils) store a large amount of carbon which is rapidly lost when these peatlands are 

drained for agriculture and commercial forestry (Hooijer et al., 2010). A rapid increase in decomposition 

rates leads to increased emissions of CO 2 

, and N 2 

O, and vulnerability to further impacts through fire. 


55

The FAO emissions database estimates globally there are 250 000 km 2 of drained organic soils under cropland 

and grassland, with total GHG emissions of 0.9 Gt CO2eq yr -1 in 2010. The largest contributions are from Asia 

(0.44 Gt CO2eq yr -1 ) and Europe (0.18 Gt CO2eq yr -1 ; FAO, 2013). Joosten (2010) estimated that there are 

>500 000 km 2 of drained peatlands in the world including under forests, with CO 2 

emissions having increased 

from 1.06 Gt CO 2 

yr -1 in 1990 to 1.30 Gt CO 2 

yr -1 in 2008. This is despite a decreasing trend in Annex I countries, 

from 0.65 to 0.49 Gt CO 2 

yr -1 , primarily due to natural and artificial rewetting of peatlands. In Southeast Asia, 

CO 2 

emissions from drained peatlands in 2006 were 0.61 ± 0.25 Gt CO 2 

yr -1 (Hooijer et al., 2010). 

Soil drainage also affects mineral soils. Meersmans et al. (2009) showed that initially poorly drained 

valley soils in Belgium have lost significant amount of topsoil SOC (e.g. between – 2 and – 4 kg C m -2 for the 

1960-2006 period). The cause is most probably intensified soil drainage in these environment for cultivation 

purposes. 

A serious consequence of deforestation is extensive loss of carbon from the soil, a process regulated by 

microbial diversity. Crowther et al. (2014) assessed the effects of deforestation on soil microbial communities 

across multiple biomes, drawing on data from eleven regions ranging from Hawaii to Northern Alaska. The 

magnitude of the vegetation effect varied between sites. Deforestation dramatically altered the microbial 

communities in sandy soils, while the effects were minimal in clay-rich soils, even after extensive tree 

removal. Fine soil particles have a larger surface area to bind nutrients and water. This capacity might buffer 

soil microbes in clay-rich soils against the disturbance of deforestation. Sandy soils, by contrast, have larger 

particles with less surface area and so retain fewer nutrients and less organic matter. Microbial community 

changes were associated with distinct changes in the microbial catabolic profile. 

Dynamic Global Vegetation Models (DGVMs) can be used to look at the combined effects of land use 

change, climate, CO 2 

, and in some cases N deposition, on vegetation and soil properties over time. In Table 

4.1, Figure 4.4 and Figure 4.5 we show results from three vegetation models: ISAM (Jain et al., 2013; El-Masri et 

al., 2013; Barman et al., 2014 a, b), LPJ-GUESS (Smith et al., 2001; Pugh et al., 2015) and LPJmL (Bondeau et al., 

2007; Schaphoff et al., 2013). The ISAM model includes a nitrogen cycle, N deposition and changes in soil N. 

The ISAM and LPJ-GUESS models were run with the HYDE historical land use change data set (History Database 

of the Global Environment, Klein Goldewijk et al., 2011). The LPJmL group combined three land use change data 

sets (Klein Goldewijk and Drecht, 2006; Ramankutty et al., 2008; Portmann, Siebert and Döll, 2010) with the 

global geographic distribution of agricultural lands in the year 2000 (Fader et al., 2010). The models were also 

run with historical climate and CO 2 

(and N deposition in the case of ISAM). Figure 4.4 shows the mineral soil C 

and N concentration of different land cover types in different geographic ranges while Table 4.1 and Figure 4.5 

show the loss of carbon due to historical land use change from 1860 to 2010. 

Differences between the models are large for some systems and regions due to different landuse change 

data, different land cover definitions, different processes included in the models, etc. For example, soil carbon 

losses are higher in the LPJmL model in part due to greater land cover change in their land cover reconstructions. 

The highest carbon losses are associated with the conversion of forests to croplands (Figures 4.4 and 4.5). 

While Table 4.1 shows the global mean soil carbon loss, the effects are not the same everywhere (Figure 4.5). 

This may be the case, for example, when forests are converted to pastures in regions where pastures strongly 

favour soil C accumulation. 


56

70 

60 

(A) soil carbon 

50 

Soil Carbon (kg/m2) 

40 

30 

20 

10 

0 

All "natural" 

forest 

grassland 

shrubland 

crop 

pasture 


forest 

grassland 

shrubland 

crop 

pasture 


forest 

grassland 

shrubland 

crop 

pasture 

ISAM 

LPJml 

LPJ_GUESS 

Tropics Temperate Boreal 

70 

60 

(A) soil carbon 

50 

Soil Carbon (kg/m2) 

40 

30 

20 

10 

0 


forest 

grassland 

shrubland 

crop 

pasture 


forest 

grassland 

shrubland 

crop 

pasture 


forest 

grassland 

shrubland 

crop 

pasture 

ISAM 

LPJml 

LPJ_GUESS 

Tropics Temperate Boreal 

Figure 4.4 Soil carbon and nitrogen under different land cover types. Source: Smith et al. (in press). 

Panel (a) shows mean soil carbon stocks; Panel (b) shows mean soil nitrogen stocks. Based on three 

vegetation models ISAM (Jain et al., 2013; El-Masri et al., 2013; Barman, Jain and Liang, 2014 a, b), LPJ-GUESS 

(Smith et al., 2001; Pugh et al., 2014); and LPJmL (Bondeau et al., 2007; Schaphoff et al., 2013). The soil carbon 

and soil nitrogen are the average over the period 2001 to 2010 (2003 for LPJmL) in model simulations with 

historical land-use change, climate, and CO 2 

(and N 2 

for the ISAM model). All ‘natural’ land is the mean of all 

lands without pasture or crop land cover. It includes ‘un-managed’ forest, grassland and shrubland categories 

and may include other land cover types depending on the models e.g. bare soil. 


57

Figure 4.5 Maps of change in soil carbon due to 

land use change and land management from 

1860 to 2010 from three vegetation models. 

Pink indicates loss of soil carbon, blue indicates 

carbon gain. The models were run with historical 

land use change. This was compared to a 

model run with only natural vegetation cover 

to diagnose the difference in soil carbon due to 

land cover change. Both model runs included 

historical climate and CO 2 

change. Source: Smith 

et al. (in press). 

Panel (a) of Figure 4.5 shows cropland and pasture coverage in 2003. The model was run with historical land 

use change. This was compared to a model run with only natural vegetation cover to diagnose the difference 

in soil carbon due to land cover change up to year 2003 as shown in Panel (b). Both model runs included 


change. Pink indicates loss of carbon due to land use, blue indicates areas of carbon 

gain. 

Model Tropical Temperate Boreal Global 

LPJ-GUESS 12.63 15.01 0.37 29.85 

LPJmL 34.86 25.99 0.05 61.86 

ISAM 17.24 37.83 5.28 60.35 

Mean 21.57666667 26.27666667 1.9 50.68666667 

Table 4.1 Soil carbon lost globally due to land use change over the period 1860 to 2010 (PgC) 

Data are from three vegetation models ISAM (Jain et al., 2013; El-Masri et al., 2013; Barman, Jain and Liang, 

2014 a, b); LPJ-GUESS (Smith et al., 2001; Pugh et al., 2015); and LPJmL (Bondeau et al., 2007; Schaphoff et al., 

2013). Each model is run with and without historical land use change data and the difference between the ‘with 

land use change’ and ‘no land use change’ runs gives the loss due to land use change. The runs also included 


and cover the period from 1900 to 2010. 

Impacts of land management and degradation 

Logging and fire are the major causes of forest degradation in the tropics (Bryan et al., 2013). Logging removes 

nutrients. Logging operations also cause soil disturbance affecting soil physical properties and nutrient levels 

(soil and litter) in tropical (e.g. Olander et al., 2005; Villela et al., 2006; Alexander, 2012) and temperate forests 

(Perez et al., 2009). Many physical, chemical, mineralogical, and biological soil properties can be affected by 

forest fires depending on fire regime (Certini, 2005). Increased frequency of fires contributes to degradation 

and reduces the resilience of the biomes to natural disturbances 


58

A meta-analysis of 57 publications (Nave et al., 2011) showed that fire had significant overall effects on soil 

C (-26 percent) and soil N (-22 percent). Fires reduced forest floor storage (pool sizes only) by an average of 59 

percent (C) and 50 percent (N), but the concentrations of these two elements did not change. Prescribed fires 

caused smaller reductions in C and N storage (-46 percent and – 35 percent) than wildfires (-67 percent and – 

69 percent). Burned forest floors recovered their C and N pools in an average of 128 and 103 years, respectively. 

Among mineral soil layers, there were no significant changes in C or N storage, but C and N concentrations 

declined significantly ( -1 1 percent and – 12 percent, respectively). Mineral soil C and N concentrations were 

significantly reduced in response to wildfires but not after prescribed burning. 

A large field study in the Amazon (225 forest plots) examined the effects of anthropogenic forest disturbance 

(selective logging, fire, and fragmentation) on soil carbon pools. Results showed that the first 30 cm of the soil 

pool did not differ between disturbed primary forests and undisturbed areas of forest, suggesting a resistance 

to impacts from selective logging and understory fires (Berenguer et al., 2014). However, impacts of human 

disturbances on the soil carbon are of particular concern in tropical forests growing on organic soils. 

Forest fires produce pyrogenic carbonaceous matter (PCM), which can contain significant amounts of fused 

aromatic pyrogenic C (often also called black C), some of which can be preserved in soils over centuries and 

even millennia. This was found to be the reason for similar soil organic C contents modelled for scenarios with 

and without burning in Australia: the loss in litter C input by fire was compensated by the greater persistence 

of the pyrogenic C (Lehmann et al., 2008). Dissolved pyrogenic carbon (DPyC) from burning of the Brazilian 

Atlantic forest continued to be mobilized from the watershed each year in the rainy season, despite the fact 

that widespread forest burning ceased in 1973 (Dittmar et al., 2012). Fire events are a source of carbonaceous 

aerosol emissions, and these are considered a major source of global warming (Kaufman, Tanre and Boucher, 

2002) 

Shifting cultivation practices of clearing land through fire have been used for thousands of years but in 

recent years increasing demographic pressure has often reduced the duration of the fallow period and so 

affected system sustainability. A review by Ribeiro Filho, Adams and Sereni Murrieta (2013) reported negative 

impact on SOC associated with the conversion stage, although impacts depended on the characteristics of 

the burning. Chop-and-mulch of enriched fallows appears to be a promising alternative to slash-and-burn. A 

study in the Amazon (Comtea et al., 2012) found that this technique conserves soil bulk density and significantly 

increases nutrient concentrations and organic matter content compared to burnt cropland and to a control 

forest. 

Climate change and land use dynamics are the major drivers of dryland degradation with important 

feedbacks through changes in plant community composition – for example shrub encroachment or decrease 

in vegetation cover (D’Odorico et al., 2013). A review conducted by Ravi et al. (2010) indicated soil erosion as 

the most widespread form of land degradation in drylands, with wind and water erosion of dryland soils 

accounting for 87 percent of the land degradation. Grazing pressure, loss of vegetation cover, and the lack 

of adequate soil conservation practices increase the susceptibility of these soils to erosion. An analysis of 

224 dryland sites highlighted a negative effect of aridity on the concentration of soil organic C and total N, 

but a positive effect on the concentration of inorganic P (Delgado-Baquerizo et al., 2013). Because aridity is 

negatively related to plant cover, the authors argue that these effects might be related to the dominance in 

arid areas of physical processes such as rock weathering, a major source of P to ecosystems, over biological 

processes that provide more C and N, such as litter decomposition. 

Grasslands, including rangelands, shrublands, pastureland, and cropland sown with pasture and fodder 

crops, covered approximately 3.5 billion ha in 2000. This represented 26 percent of the global ice-free land 

area and 70 percent of the agricultural area, and contained about 20 percent of the world’s soil organic carbon 

(C) stocks. Portions of the grasslands on every continent have been degraded due to human activities – about 


59

7.5 percent of grassland worldwide has been degraded because of overgrazing (Conant, 2012). Grassland 

management and grazing intensity can affect the stock of SOC. A multifactorial meta-analysis of grazer 

effects on SOC density (17 studies that include grazed and ungrazed plots) found a significant interaction 

between grazing intensity and grass type. Specifically, higher grazing intensity was associated with increased 

SOC in grasslands dominated by C 4 

grasses (increase of SOC by 6–7 percent), but with lower SOC in grasslands 

dominated by C 3 

grasses (decrease of SOC by an average 18 percent). Impacts of grazing were also influenced 

by precipitation. An increase in mean annual precipitation of 600 mm resulted in a 24 percent decrease in 

grazer effect size on finer textured soils, while on sandy soils the same increase in precipitation produced a 22 

percent increase in grazer effect on SOC (McSherry and Ritchie, 2013). 

4.3.2 | Land use intensity change 

Land use intensity has increased in recent decades, largely driven by the need to feed a growing population, 

by shifts in dietary patterns towards more meat consumption, and by the growing production of biofuels. At 

the same time, fast urbanization has occupied more of the land, reducing the stock available for agricultural 

production. Intensification has been widely advocated because of the many negative environmental 

consequences of clearing natural ecosystems to expand agricultural areas. 

However, intensifying management practices, such as fertilization, irrigation, tillage and increased 

livestock density, can have negative environmental impacts (Tilman et al., 2002). Intensifying land use can 

potentially reduce soil fertility. Intensification can also reduce soil resilience to extreme weather under climate 

change, to pests and biological invasion, to environmental pollutants and to other disasters. This section 

provides an overview of the benefits and consequences of intensifying use of agricultural lands. The section 

also highlights examples of how negative consequences can be minimized. 

Several factors influence the increase in land use intensity during the recent decades. On the demand side, 

three main factors are at play: (i) the need to meet the food, fibre, and fuel demands of a growing population; 

(ii) an increase in meat consumption as developing nations become wealthier and tastes change; and (iii) 

rising demand for crops for biofuels. On the supply side, settlements are occupying more land and so reducing 

the land available for agriculture. 

To meet the increased demand, it is estimated that food production will need to increase by 70 -1 00 percent 

by 2050 (World Bank, 2008; Royal Society of London, 2009; Keating et al., 2014). Of the two pathways of 

increasing production—intensification and expansion—intensification is widely promoted as the more 

sustainable option because of the negative environmental consequences of land expansion through 

deforestation and conversion of wetlands to cultivation (Foley et al., 2011; MA, 2005). However, the current 

increase in land use intensity is generally not sustainable. In order to give a clear picture of the effects of 

increased land use intensity, this section is organized according to the primary management practices that 

characterize intensification of agricultural lands (see Table 4.2 for summary). 

Nutrient management 

Nutrient inputs, from both natural and synthetic sources, are needed to sustain soil fertility and to supply 

the nutrient needs of higher yielding crop production. Intensification in recent years has led to the annual 

global flows of nitrogen and phosphorus now being more than double the natural levels (Matson et al., 

1997; Smil, 2000; Tilman, 2002). The trend is still increasing – in China, for example, N input in agriculture in 

the 2000s was more than double the levels of the 1980s (State Bureau of Statistics-China, 2005). Nutrient 

management is particularly intensive in greenhouse production systems. In some parts of Asia, for example, 

up to six tons of chemical nutrient and hundreds tons of organic fertilizers are applied per hectare each year 

in order to achieve high yielding multiple cropping of vegetables (Liu et al., 2008). Between 50-60 percent of 


60

the nutrient inputs remain in the croplands after harvest (West et al., 2014). When these nutrients are later 

mobilized, they become a major source of pollution to local, regional and coastal waters (Carpenter et al., 

1998). Intensive nutrient input in agriculture has been shown to be a major cause of eutrophication and algae 

blooming in lakes and inshore waters. In addition, over-use of nitrogen chemical fertilizers has been found in 

many locations globally to be a cause of acidification and accelerated decomposition of soil organic matter, 

leading to further soil degradation in over-fertilized soils (Ju et al., 2009; Tian et al., 2012). 

Nutrient inputs also affect the earth’s climate. Globally, approximately one percent of nitrogen additions 

are released to the atmosphere as nitrous oxide (N 2 

O), a gas which has 300 times the warming power of 

carbon dioxide (Klein Goldewijk and van Drecht, 2006). China, India, and the United States account for ~56 

percent of all N 2 

O emissions from croplands, with 28 percent originating from China alone (West et al., 2014). 

One remedy is to increase the efficiency of nutrient use. Nutrient efficiency can be significantly increased – 

and N 2 

O emissions can be reduced – through changes in the rate, timing, placement, and type of application 

of nutrients, and by improving the balance amongst nutrients applied (Venterea et al., 2011). In addition, if best 

management practices are used, agricultural soils have the potential to be carbon storage areas (Paustian 

et al., 2004; Smith, 2004). Technological improvements are being made to the production of biochar which 

converts a fraction of the C present in the original material into a more persistent form through carbonisation. 

Biochar can then be used as a soil amendment to provide agronomic and environmental benefits (Lehmann 

and Joseph, 2015). In many cases, the presence of biochar has caused a reduction in N 2 

O emissions, especially 

when these originate from denitrification. However, the mechanics of the process are not yet fully understood 

(Cayuela et al., 2013; 2014). 

The effect of pesticides on soil biodiversity 

The large-scale use of pesticides may have direct or indirect effects on soil biodiversity. With the 

intensification of agriculture, the use of pesticides has increased worldwide to approximately two million 

tonnes per year (herbicides 47.5 percent, insecticides 29.5 percent, fungicides 17.5 percent, other 5.5 percent 

by De et al. (2014)). Studies of the effect that pesticides have on soil biodiversity have shown contradictory 

results. Effects are dependent on a variety of factors including the chemical composition, the rate applied, 

the buffering capacity of the soil, the soil organisms in question, and the time-scale. For example, Boldt and 

Jacobsen (1998) tested the effects of sulfonylurea herbicides on strains of fluorescent pseudomonads cultured 

from agricultural field soils. They found that the herbicide Metsulfuron methyl was toxic to the majority of 

fluorescent pseaudomonads (77 strains) in low concentrations, while Chlorsulfuron was only toxic at high 

concentrations, and Thifensulfuron methyl was toxic only to a few strains, even at high concentrations. 

In a review by Bünemann, Schwenke and Van Zwieten (2006) of the effects of pesticide application on 

soil organisms, there were no data available for 325 of 380 active constituent pesticides registered for use in 

Australia. The review thus effectively highlighted the huge gap in knowledge. A synthesis of the impact of 

herbicides on non-target organisms concluded that herbicides did not have a major effect on soil organisms 

(Bünemann, Schwenke and Van Zwieten, 2006) with the exception of butachlor, which was toxic to 

earthworms when applied at typical agricultural rates (Panda and Sahu, 2004). In addition, the application of 

bromoxynil herbicides caused a shift in the communities of four out of five targeted bacterial taxa even after 

degradation of the herbicide (Baxter and Cummings, 2008). Avoidance behaviour to phendimedipham has 

also been observed for collembola (Heupel, 2002) and earthworms (Amorim, Rombke and Soares, 2005). 

Insecticide application, however, has a much greater effect on soil biota, including changes in microbial 

community composition (Pandey and Singh, 2004), lower collembolan abundance (Endlweber, Schadler and 

Scheu, 2005) and earthworm reproduction. Because some species of earthworm such as Eisenia Fetida can be 

easily bred and because they ingest large quantities of organic matter in the soil, earthworms have often been 

used as bioindicators of chemical toxicity in soils (Yasmin and D’Souza, 2010). A variety of studies have reported 

changes in earthworm reproductive rates, growth rates and weight loss when the pesticides Malathion 


61

(Espinoza-Navarro and Bustos-Obregon, 2005), Chlorpyrifos (Zhou et al., 2007; De Silva et al., 2010), Benomyl 

(Römbke, Garcia and Scheffczyk, 2007), Carbofuran (De Silva et al., 2010) were applied to soil in laboratory 

experiments. Non-target effects of insecticide applications may be highly dependent on the organism since 

field application of Chlorpyrifos did not affect the abundance of soil predatory mites (Navarro-Campos et al., 

2012). Fungicides have also demonstrated significant negative effects on earthworms (Eijsackers et al., 2005). 

In particular, copper-based fungicides that are resistant to degradation have caused long-term reductions in 

earthworm populations (Van Zwieten et al., 2004). 

Although an assessment of soil food webs across Europe did not specifically focus on pesticide application, 

the study demonstrated that land-use intensification was related to decreased diversity of soil fauna and 

resulted in less diversity among functional groups. Larger soil animals showed the most sensitivity (Tsiafouli 

et al., 2015). However, there have been no such comprehensive studies to quantify the effects of pesticides on 

soil organisms at multiple trophic levels across regions. Such studies need to consider also the indirect effect 

of pesticides, including interactions between pesticides and biotic factors. Since below-ground biodiversity 

is intimately linked to above-ground vegetation patterns (De Deyn and van der Putten, 2005) and vice versa 

(Bardgett and van der Putten, 2014), changes in plant diversity resulting from herbicide may cause indirect 

effects of herbicide application. 

Water management 

The area of irrigated croplands has doubled in the last 50 years and irrigation now accounts for 70 percent 

of all water diversions on the planet (Gleick, 2003). Irrigated areas account for 34 percent of crop production, 

yet only cover 24 percent of all cropland area (Siebert and Doll, 2010). With the increased frequency of 

drought under climate change, demand for agricultural water is rising in many locations. Not surprisingly, 

irrigation is most commonly used in more arid areas. Where a high proportion of available water is used for 

agriculture, this can cause water stress for both people and nature. Water efficiency can be improved through 

infrastructure and through better management practices. Irrigation can potentially increase soil salinity in 

dry regions (Ghassemi, Jakeman and Nix, 1995). Where salinization occurs, additional irrigation is needed to 

‘flush’ the salts beyond the root zone of the crops. This additional water requirement can further exacerbate 

water stress. 

Harvest frequency 

Land use intensity can also be increased by harvesting a parcel of farmland more frequently (double 

cropping, triple cropping). Approximately 9 percent of crop production increases from 1961-2007 came from 

increases in the harvest frequency (Alexandratos and Bruinsma, 2012). As more land was double cropped, the 

global harvested area increased four times faster than total cropland between 2000 and 2011 (Ray and Foley, 

2013). In addition, with global warming, the areas suited for double or even triple cropping are extending into 

subtropical and warm temperate regions (Liu et al., 2013a). The factors involved in this fast rate of increase 

include: fewer crop failures; fewer fallow years; and an increase in multi-cropping. 

Greenhouse production has allowed multiple cropping around the world. For fruit and vegetable crops, 

world greenhouse cultivated area reached a total area of 408 890 ha in 2013, which includes as many as five 

harvests in a single year. This increasing harvest frequency has reduced soil quality through soil compaction 

and has increased the risk of pathogen diseases. The intensive use of pesticides and herbicides in greenhouses 

not only affects soil quality but creates risks to human health. In some greenhouse systems, long term 

multiple cropping has led to soil acidification, salinization and biological deterioration, especially where large 

amounts of fertilizer and pesticide/herbicide have been used. In these situations, there is a need to improve 

management practices, using organic matter, balancing nutrient additions and adopting intermittent fallow. 


62

Livestock density 

Livestock production is projected to increase to meet the growing demand for livestock products from 

a rising population and from an increase in per capita consumption. The greatest increases in per capita 

demand are projected to be in developing and transition countries (Bouwman et al., 2006). Since the 1970s, 

most increases in livestock production have resulted from intensification, with a shift to a greater fraction 

of livestock raised in industrial conditions (Bouwmann et al., 2006). For example, 76-79 percent of pork and 

poultry production is now industrialized (Herrero et al., 2013). 

Industrial livestock production systems can be highly polluting. The manure from animals, the inputs for 

growing animal feed, and the soil loss from intensively managed areas can all be major sources of water 

pollution to local and downstream freshwater ecosystems. Where natural ecosystems are cleared and 

converted to pasture, particularly in arid and semi-arid regions, the lands are typically low potential and have 

a high risk of soil erosion and soil carbon/nutrient depletion (Delgado et al., 1999; Seré and Steinfeld, 1996). 

The soils capacity for water storage and their biodiversity are also at risk. Moreover, intensified livestock 

production requires an increased use of veterinary medicines, sulfa-antibiotics and hormones, all of which 

carry risks of pollution to soil, water and the livestock products themselves, with risks to biological and human 

health. 

Forestry harvest and wetland draining 

Forests and wetlands and their soils are massive reservoirs of carbon. In fact, forest soils store approximately 

the same amount of carbon as the living biomass of the forest itself (FAO, 2010). Wetlands are important 

not only for the huge carbon pool they contain but also for their role in the hydrological cycle. However, 

wetlands along big river banks, lakes and estuaries have been increasingly developed for croplands/bioenergy 

production in recent decades, particularly in Asia. The majority of soil carbon is concentrated in peatlands 

within the boreal forest as well as tropical forests in Southeast Asia. Around the world, deforestation causes 

~25 percent of the total loss of soil carbon (Guo and Gifford, 2002; Murty et al., 2002). This loss largely stems 

from oxidation of the organic matter and from soil erosion. In China over the last four decades, almost 1.3 

million ha of wetlands have been converted to crop production, causing the loss of about 1.5 Pg C of soil carbon 

(Zhang et al., 2009). Deforestation continues through conversion to agriculture and through extraction of 

forest products. Between 2000 and 2012, there was a new loss of 1.5 million square kilometres of forests, with 

the most pronounced trend in the tropics (Hansen et al. 2013). Soil erosion and organic matter oxidation can 

be reduced through selective tree harvesting rather than clear felling, and by avoiding deforestation on steep 

slopes. Draining and cultivating wetlands can also affect local and regional water storage. 


63

Research needs 

It will be evident from the discussion in this section that much remains to be learned. Amongst the priority 

research questions are the following: 

1. Sustainable intensification – How can we get the benefits from intensification while minimizing the 

associated environmental and social costs? 

2. Trade-offs between soils and efficiency – How can we manage for resilient soil and related ecosystem 

services while continuing to maximize efficiency? To what extent can we have both? 

3. Soil degradation and intensification – What is the extent of degraded soils? There are currently no sound 

estimates. What portion of degraded soils can be attributed to un-sustainable intensification? 

4. Options and trade-offs for improved soil management – What can we learn from management practices 

used in intensification areas to help restore degraded soils? Are there any options that can integrate 

best management practice for sustainable intensification? What are the short – and long-term tradeoffs 

of resource use and sustainability? What are the environmental and social costs and economic 

benefits of land use intensification? 

5. Farming practices and soil health – How do changes in harvest frequency and crop rotation affect soil 

resilience? How much change is needed to restore degraded soils? 

Land 

intensification 

Sector Distribution Major 

environmental 

consequence 

Knowledge gap 

Cropping 


Harvest 

frequency 

Globally 

Soil quality and 

resilience 

Ecosystem service 

Continuing 

monoculture 

Developing 

and transition 

countries 

Soil health, 

pesticide residue 

Biological resilience 

Nutrient 


Over fertilization 

Developing 

countries 

Soil acidification, 

water pollution, 

N 2 

O emission 

and nitrate 

accumulation 

Rate reducing versus 

balancing? 

Irrigation Submerged Rice Developing 

countries, Asia 

Water scarcity, 

methane 

emission 

Trade-offs C and water, 

Dry crops 

Arid/semi-arid 

regions 

Secondary 

salinization, 

water scarcity 

Competition for water 

Livestock 


Over grazing 

Developing 

countries 

Soil degradation, 

water storage, 

C loss 

Forage versus feed crops? 

Industrial 

breeding 

Industrialized 

countries 

Waste, water 

pollution, residue 

of veterinary 

medicine and 

antibiotics 

Safe waste treatment 

and recycling 

Forest clearance, 

wetlands 

drainage 

Deforestation. 

wetland shrink 

Developing 

and transition 

countries 

Biodiversity, 

natural wealth, 

C loss 

Agro-benefit versus 

natural value 

Table 4.2 Threats to soil resource quality and functioning under agricultural intensification 


64

4.3.3 | Land use change resulting in irreversible soil change 

In this section we deal with soil sealing and mining, which have been identified as two important soil 

degradation processes occurring around the world. The current extent and rate of growth of soil sealing and 

mining are significant, and create considerable risks to essential ecosystem services. These changes in land 

use nearly always require a trade-off between various social, economic and environmental needs. 

Sealing and land take 

The ongoing urbanization and conversion of the landscape with settlements, infrastructure and services is 

occurring in many regions. Europe and Asia, in particular, are experiencing high rates of urban expansion and 

urban sprawl, and there are often insufficient incentives to re-use brownfield sites. These factors are causing 

an increase in land take and soil sealing. The drivers are essentially economic and demographic growth. In 

Europe, America and Oceania, at least 70 to 80 percent of the population currently lives in urban areas. The 

rate of urbanization is expected to continue to increase, particularly in Asia and Africa. 

The concept of land take covers all forms of conversion for the purpose of settlement, including: the 

development of scattered settlements in rural areas; the expansion of urban areas around an urban nucleus; 

the conversion of land within an urban area (densification); and the expansion of transport infrastructure such 

as roads, highways and railways. Broadly, this discussion considers as land take any conversion of agricultural, 

natural or semi-natural land cover to an ‘artificial’ (e.g. human-made) area. Artificial land cover classes are 

categorized in the Corine Land Cover system – see Table 4.3. 

A greater or smaller part of land take will result in soil sealing. Soil sealing means the permanent covering 

of an area of land and its soil by impermeable artificial material such as asphalt or concrete, for example 

through buildings and roads. As shown in Figure 4.6, the sealed area is only part of a settlement area. Gardens, 

urban parks, leisure areas and other green spaces within the boundaries of settlements are not covered by an 

impervious surface or are only partially covered. They thus form part of a land take but do not contribute to 

soil sealing (Prokop, Jobstmann and Schöbauer, 2011.) The ratio between sealed area and total area for a given 

land use class is measured by the soil sealing index. An example of this index, calculated for the Italian region 

of Emilia-Romagna, is shown in Table 4.4. 

Table 4.3 Artificial areas in Corine Land Cover Legend 


65

A) Typical structure of settlement B) Sealed areas about 70 percent (black color) 

Figure 4.6 Schematic diagram showing areas sealed (B) as a result of infrastructure development for a settlement (A). Source: 

European Union, 2012. 

Table 4.4 Artificial areas in Emilia Romagna according to the Corine Land Cover Legend and 

sealing index. 


66

Impact of land take 

Land take, by its definition, is the subtraction of an area from a previous agricultural, natural or semi-natural 

land use. According to this definition, the most obvious impact on the ecosystem services that can be provided 

by soil is on the production of biomass, and in particular of food. To clarify the concept, we may imagine that 

a city expands its urbanized area by a new allotment of 100 ha created at the expense of agricultural land. This 

area will be covered by buildings, private and public gardens, commercial centres, roads, etc. The entire area 

will clearly lose most of its capacity to produce food, with the possible minor exception of family horticulture 

in unsealed areas such as gardens or allotments. Had the entire area been previously cultivated with, say, 

winter wheat with an average yield of 5 tonnes ha -1 , the total loss in terms of food production potential will be 

equal to 500 tonnes of winter wheat per year. 

Other ecosystem services are at risk also. Water infiltration and purification and carbon storage are mainly 

reduced by the effective sealed area, and not by the entire land taken. Support to biodiversity is clearly affected, 

although the degree depends on the different groups of organisms and also on the design of the urbanized area. 

In this context, a positive mitigation role can be played by ‘Urban Green Infrastructure’ – the incorporation of 

a network of high-quality green spaces and other environmental features. Green Infrastructure can include 

natural areas as well as human-made rural and urban elements such as urban green spaces, reforestation 

zones, green bridges, green roofs, eco-ducts to allow crossing of linear barriers, corridors, parks, restored 

floodplains, biodiverse farmland. 

Regulation of land take and mitigation of its impacts 

Where policy aims to minimize land take, measures can be implemented to encourage re-use of existing 

urban areas such as derelict areas, brownfields and upgrading of degraded neighborhoods. Measures 

promoting densification of existing urban areas can also contribute to the reduction of land take. 

Fiscal measures can prevent speculative urban sprawl. A number of municipalities, and regional 

governments, especially in Europe, have already adopted policies designed to achieve zero net urban expansion. 

However, zero expansion becomes more problematic when there is significant demographic pressure and a 

high rate of rural to urban migration. 

Rational and efficient urban planning and intelligent building and infrastructure design can also help 

reduce land take. In the past, urban planners, architects and civil engineers too often considered soil as a raw 

material, abundantly available and of limited value. Examples of efficient consideration of the value of soil in 

urban development include: the construction of parking lots in the basement of buildings; and ‘green’ covering 

of areas that are only occasionally used, such as parking lots for exhibitions and fairs etc. 

Where expansion of urban and built-up areas is a policy and planning imperative, intelligent urban planning 

needs to take account of the soil dimension to mitigate the impact of land take. An education process is 

needed to make urban planners aware of the value of soil quality and land capability and of the options for 

mitigating negative impacts of land take. 

Impacts of soil sealing 

Sealing by its nature has a major effect on soil, diminishing many of its benefits. Normal construction 

practice is to remove the upper layer of topsoil, which delivers most of the soil-related ecosystem services, 

in order to be able to develop strong foundations in the subsoil or underlying rock to support the building or 

infrastructure. Where strong foundations are not required, only a thin layer of topsoil is generally excavated 

and the surfaces are simply covered by a layer of impervious material, such as asphalt or concrete. Both 

techniques impair or eliminate the soil’s capacity to deliver ecosystem services. 


67

The main impacts include the following. 

1. Water infiltration and purification are lost, and regulation of the water cycle is completely altered. The 

concentration time of water flow is shortened, promoting flood events. 

2. Soil biodiversity is impaired, as sealing prevents the production, release and recycling of organic material, 

so affecting the soil biological communities (Marfenina et al., 2008). In addition, the alteration of soil 

water regimes, soil structure and redox potential have a strong impact on soil biodiversity. 

3. Soil carbon storage potential is fundamentally altered (Jones et al., 2005), particularly where topsoil, 

which normally contains about half of the organic carbon in mineral soils, is stripped off. 

4. The urban microclimate is altered. The reduction of evapotranspiration in urban areas due to the loss of 

vegetation and through alteration of albedo strengthens the ‘urban heat island’ effect (Früh et al., 2011). 

Prevention of soil sealing and mitigation of its impacts 

Appropriate mitigation measures can be taken in order to maintain some of the ecosystem functions of 

soils and to reduce negative effects on the environment and human well-being. Key options available to urban 

planners and managers include: (i) minimizing conversion of green areas; (ii) re-use of already built-up areas, 

such as brownfield sites; (iii) using permeable cover materials instead of concrete or asphalt; (iv) supporting 

Green Infrastructure (see above); and (v) providing incentives to developers to minimize soil sealing. 

In practice, planners need to be able to evaluate the tradeoffs and ensure that policy instruments are used 

to ensure optimal outcomes which consider both human needs for urbanization and the preservation of the 

integrity of the soil and its services: 

1. Existing policies for development of settlements and infrastructure should be reviewed and adapted to 

take account of the value of soils, particularly where subsidies or other incentives are driving unplanned 

land take and soil sealing (Prokop, Jobstmann and Schöbauer, 2011). 

2. An integrated approach to urban planning should be followed. Existing best practice has demonstrated 

that soil sealing can be limited, mitigated and compensated. This requires that spatial planning 

follow an integrated approach and involve the full commitment of all relevant public authorities and 

governance entities responsible for land management, such as municipalities, counties and regions 

(Siebielec et al., 2010). 

3. Specific regional and local approaches can be developed. These could, for example, take into account 

unused resources at the local level such as a particularly large number of empty buildings or brownfield 

sites. 

Mining 

Ancient mining 

Mining is the extraction from the Earth of rocks, valuable minerals, and other geological materials of economic 

interest. It is one of the most ancient activities in human history (Mighall et al., 2002; Shotyk et al., 1998). 

Mining for specific materials such as quartz, silex and clays began as far back as the Palaeolithic – the Old 

Stone Age – when the first stone tools were developed. In the Neolithic era – the New Stone Age – flint mines 

existed in Belgium, Britain and elsewhere. Landscape records and evidence from bogs show that mining 

activities became more intense with the development of metal tools in the Bronze Age, and subsequently 

in the Iron Age (Martínez-Cortizas et al., 2002; Shotyk et al., 1998). Examples of the environmental impact 

of ancient mining are numerous (Figure 4.7) (López – Merino et al., 2010; Grattan, Huxley and Pyatt, 2003; 

Fernández Caliani, 2008). 


68

Photo by F. Macias 

Photo by J.C. Fernández Caliani 

Photo by F. Macias 


Figure 4.7 (A) Panoramic view of Las Medulas opencast gold mine (NW Spain). The Roman extractive technique – known as ‘ruina 

montis’ – involved the massive use of water that resulted in important geomorphological changes; (B) Weathered gossan of the Rio 

Tinto Cu mine, considered the birthplace of the Copper and Bronze Ages; (C) typical colour of Rio Tinto (‘red river’ in Spanish), one of 

the best known examples of formation of acid mine waters. These are inhabited by extremophile organisms. 

Impact of mining 

The impact of mining on the environment differs greatly depending on the type of extraction, the ore 

or material exploited, and the method used to process the material extracted (Moore and Luoma, 1990). 

Traditional underground mining, which follows profitable veins beneath the earth’s surface, has less impact 

than open cast mining activities – also referred to as strip mining − which grew very rapidly in the last hundred 

years (Salomons, 1995). In some instances, entire mountains have been literally blasted apart to reach thin ore 

vein seams within, leaving permanent scars on the landscape. Nonetheless, mining operations themselves 

affect relatively small areas. By contrast, significant environmental problems are caused by tailing and waste 

rock deposits and by subsequent smelting operations. Pollutants can be transferred to surrounding areas by 

acid mine drainage or by atmospheric deposition of wind-blown dust. The incidence of these problems depends 

on local climatic and hydrologic conditions (Aslibekian and Moles 2003; Batista, Abreu and Serrano, 2007; 

López, Gónzalez and Romero, 2008). Other environmental effects, in addition to those caused by pollutants, 

include deforestation, erosion and formation of sinkholes (Meuser, 2010; Hester and Harrison, 2001). 

Only a small fraction of the material extracted is valuable ore. The ore needs to be separated by milling 

and flotation from the large volume of other material discarded as tailings. When the remaining concentrate 

is refined by processes such as smelting, flue dust and slag are produced (Hutchinson, 1979). Atmospheric 

contamination has commonly occurred throughout the world during smelting operations, leading to 

contaminated soils and risks to livestock (Down and Stocks, 1977; Munshower, 1977). Mining for coal, gold, 

uranium, wolfram, tin, platinoids and, in particular, poly-metallic sulphides has created large environmental 

impacts on soil, water and biota. Sulphide minerals include iron sulphides such as pyrite and pyrrhothite, 

and other poly-metallic sulphides, such as those containing Cu, Pb, Zn, Hg, Cd, Tl, Sb, Bi etc. These sulphides 

can also in some instances combine with arsenides or selenides to form sulfoarsenides or sulfoselenides 

(Evangelou, 1995; Abreu et al., 2010). 

Sulphide minerals oxidise when brought to surface conditions (Nordstrom and Southam, 1997; Nordstrom 

and Alpers, 1999). The sulphide oxidation can cause extreme changes in Eh and pH (Figure 4.8) – negative 

pH values (as low as -3.6) have been measured in the acid mine waters of the Richmond mine in California 

(Nordstrom and Alpers, 1999). Depending on the local geochemical and hydrological conditions, sulphide 

oxidation can also affect the electrical conductivity of the system and may lead to elevated concentrations 

of many toxic elements in soils and waters nearby. Waters downstream of these mine systems (Figure 4.7C) 

are frequently hyperacid, hyperoxidant and hyperconductive. These waters may exhibit high activities of: (i) 

various metal species such as Al +3 , Al-SO 4 

); (ii) heavy metal species, for example Cu +2 , Cd +2 , Zn +2 , Hg +2 y Hg 0 ; 

and (iii) metalloids, including arseniates, arsenites and seleniates (Sengupta, 1993; Macías, 1996; Monterroso, 

Alvarez and Macías, 1994; Monterroso et al., 1998, 1999; Azcue, 1999). Smelting operations of sulphide minerals 

also generate SO 2 

, which, if not recovered, is released into the atmosphere and thus contributes to acid 

deposition (described in Section 4.4). 


69

The mining of gold deserves special attention given its contribution to Hg emissions (Drude de Lacerda, 

2003). Mercury is used to concentrate the fine gold particles through amalgamation and then the gold is 

separated from the amalgam by applying heat. When this process is carried out under uncontrolled conditions 

– as in small-scale gold mining (Drude de Lacerda, 2003) – Hg volatilises to the atmosphere. Tailings from Hg 

amalgamation are then leached with cyanide, and waste contaminated with metals and cyanide is released 

into the environment (Veiga et al., 2009). Arsenic exposure has also been recorded in many gold and base 

metal producing countries (Williams, 2001). However, arsenate and arsenite mobilisation can be controlled 

with soil colloidal compounds such as reactive Fe and Al (Goldberg, 2002). 

As materials from mining are exposed to the environmental conditions of the Earth’s surface, these minesoils 

develop through weathering (Sencindiver and Ammons, 2000). However, their properties differ considerably 

from the original soil. They contain a high percentage of rock fragments, a low nutrient content, and elevated 

levels of potentially harmful trace elements. They also usually lack a distinct horizonation. These soils are 

in fact very young soils characterised by properties that limit their functions and their capability to support 

vegetation (Macias, 1996; Vega et al., 2004; Abreu and Magalhães, 2009). When the overburden contains 

sulphidic material such as pyritic mine waste, the major weathering process is the oxidative dissolution 

of pyrite. Here the rate of soil formation is mainly controlled by the sulphide content and its particle-size 

distribution, causing strongly acidic conditions, as described above (Neel et al., 2003; Haering, Daniels and 

Galbraith, 2004). Quite often, restoration of mine soils requires the addition of exogenous material to correct 

the extreme pH, Eh and/or EC values and the anomalous concentrations of toxic elements common in these 

systems which are generally bioavailable and susceptible to mobilisation. 

1.2 

20 

PO2 =1 P h2 

=1 

0.6 

Hyperacid 

soils 

Circum-neutral 

soils 

Calcareous 

soils 

10 

Eh 

(volts) 

0 

-0.6 

Acid 

soils 

Thionic 

soils 

O 2 

pe 

Alkaline 

soils 

H 2 

-10 

0 

Hydric 

soils 

0 2 4 6 8 10 12 14 

pH 

Figure 4.8 Eh-pH conditions of thionic/sulfidic soils and of hyperacid soils. Source: Otero et al., 2008. 


70

The formation of sulfidic material requires strongly reducing conditions and slight acidity. Once these 

are oxidised, and in the absence of minerals with high acid buffering capacity, extremely acid and oxidising 

conditions are generated. The dashed envelope in Figure 4.8 is the approximate extent of redox-pH conditions 

of mineral soils (with the exception of hyper-acid soils). 

Preventing impacts from mining 

The rehabilitation of abandoned mines is a difficult and costly task. In fact, in many instances, the 

landscape cannot be repaired. Some mining methods may have significant environmental and public health 

effects. The Aznalcollar pyritic sludge spill (SW Spain) (López-Pamo et al., 1999; Grimalt, Ferrer and McPherson, 

1999; Aguilar et al., 2004; Calvo de Anta and Macias, 2009) is such an example. It occurred in 1998 in the 

surroundings of Doñana Park − the largest reserve of bird species in Europe − as a result of the failure of a 

tailings dam which contained several million tons of pyrite stockpile, flotation tailings and acid waters. The 

toxic spill contaminated ca. 26 km 2 of riverbanks and adjacent farmlands, extending 45 km downstream, with 

an estimated quantity of 16 000 tonnes of Zn and Pb, 10 000 tonnes of As, 4 000 tonnes of Cu, 1 000 tonnes 

of Sb, 120 tonnes of Co, 100 tonnes of Tl and Bi, 50 tonnes of Cd and Ag, 30 tonnes of Hg, and 20 tonnes of Se. 

Mining operations have a responsibility to protect the environment: air, water, soils, ecosystems and 

landscape. Many countries require reclamation plans for mining sites to follow environmental and rehabilitation 

codes. Nonetheless, mine restoration is still problematic, mainly because the environmental impacts were 

only recently understood or appreciated (Azcue, 1999; Sengupta, 1993). In addition, the technology available 

has not always been adequate to prevent or control environmental damage. Restoration of such systems 

requires a thorough understanding of material properties and their geochemistry. Only through such an 

understanding can the current and future behaviour of such systems be predicted and appropriate decisions 

taken to ensure their restoration (Gil et al., 1990; Macías-García, Camps Arbestain and Macías, 2009; Macías- 

García et al., 2009). 

Development of tailor-made Technosols to restore mine soils 

Technosols are defined by the FAO (2014b) as those soils with recent human activities in industrial and urban 

environments which have resulted in the presence of artificial and human-made objects. Technosols often 

result from the abandonment of urban, mining or industrial waste. These soils tend to have a large content of 

artefacts – that is objects that are either human-made, strongly transformed by human activity, or excavated 

(e.g. mine spoils, rubbles, cinders) (FAO, 2014b). 

Throughout history, humans have formed soils – ‘anthropogenic soils’ - and in certain cases these soils have 

proved more fertile that natural soils nearby (Sombroek, Nachtergaele and Hebel, 1993). Thus, it is feasible 

to produce specific Technosols which can fulfil the environmental and productive functions of natural soils – 

essentially, tailor-made Technosols. This may require the formulation and mixing of artefacts and other waste 

materials such as manure and biosolids. The production of these Technosols could be a feasible technique 

through which waste products are reused and the elements they contain are returned to their biogeochemical 

cycles, while restoring degraded areas and contributing to the sequestration of C in soils and biomass (Macias 

and Camps Arbestain, 2010). 

Environmental problems associated with this use of Technosols may be prevented if: (i) the characteristics 

of the materials used provide the soil with adequate buffering properties against contaminants, pH and/or 

redox changes; and (ii) there is a good understanding of how the constituent mixtures will evolve over time 

under the pedoclimatic conditions of the area to be restored. Figure 4.9 illustrates the benefits of the use 

of tailor-made Technosols in the restoration of an abandoned Cu mine rich in pyrite (Macías-García, Camps 

Arbestain and Macías, 2009; Macías-García et al., 2009; Macías and Camps Arbestain, 2010). 


71


Figure 4.9 Use of different Technosols derived from wastes in the recovery of hyperacid soils and waters in the restored mine of Touro 

(Galicia, NW Spain). 

4.4 | Atmospheric deposition 

4.4.1 | Atmospheric deposition 

The impacts of the deposition of atmospheric pollutants on soils vary with respect to soil sensitivity to 

a specific pollutant and to the total pollutant load. Anthropogenic emissions of sulphur, nitrogen and trace 

elements to the atmosphere mainly derive from fossil fuel and waste combustion in, for example, power 

generation, incineration, industry and transport. Emissions may also derive from non-combustion processes 

such as agricultural fertilizers or waste amendments. Mining activities may also contribute, for example 

Hg mining. Once in the atmosphere, these pollutants can be transported off-site and even cross national 

borders before being deposited either as dry or wet deposition. Deposition is more accentuated in forests, 

especially in coniferous forests (because of reduced wind speeds) and in areas of high elevation because of high 

precipitation rates. 

Once in the soil, pollutants can be mobilised by being: (i) released back to the atmosphere; (ii) made available 

to biota; (iii) leached out to surface waters; or (iv) transported to other areas by soil erosion. Pollutants disrupt 

natural biogeochemical cycles by altering soil functions. This disruption may come about through direct 

changes to the nutrient status, acidity, and bioavailability of toxic substances, or through indirect changes 

to soil biodiversity, plant uptake and litter inputs. Soil sensitivity to atmospheric pollution varies with respect 

to: (i) key properties influenced by geology and associated pedogenesis such as cation exchange capacity, 

soil base saturation, aluminium, or rate of base cation supply by mineral weathering); (ii) organic matter 

content and carbon to nitrogen ratio (C:N); and (iii) position of the water table. When atmospheric pollution 

is associated with sulphate deposition, the capacity of soils to adsorb sulphate (e.g. soils with a dominance 

of short-range ordered constituents) plays a key role in buffering the acidification process (Camps Arbestain, 

Barreal and Macías, 1999; Rodríguez-Lado, Montanarella and Macías, 2007). Harmful effects on soil function 

and structure occur where deposition exceeds the ‘critical load’ - the specific amount of one or more pollutants 

that a particular soil can buffer (Nilsson and Grennfelt, 1988). Estimates and mapping of critical loads of acidity 


72

are however strongly dependent on the neutralisation mechanisms considered in the analysis, for example, 

the inclusion or exclusion of sulphate adsorption (Rodríguez-Lado, Montanarella and Macías, 2007). Spatial 

differences in soil sensitivity − commonly defined by the ‘critical load’ − and in pollutant deposition result in 

an uneven global distribution of impacted soils (Figure 4.10). For instance, global emissions of sulphur and 

nitrogen have increased 3–10 fold since the pre-industrial period (van Aardenne et al., 2001), yet critical loads 

for acidification are only exceeded in 7–17 percent of the global natural terrestrial ecosystems area (Bouwman 

et al., 2002). 

4.4.2 | Main atmospheric pollutants: Synopsis of current state of knowledge 

Since the 1980s, emissions of pollutants, notably sulphur, across Europe and North America have declined. 

The decline is due to the establishment of protocols under the 1979 Convention on Long-range Transboundary 

Air Pollution (LRTAP) and the 1990 United States Clean Air Act Amendments (CAAA) (Greaver et al., 2012; Reis et 

al., 2012; EEA, 2014). Conversely, emissions in South and East Asia, sub-Saharan Africa and South America are 

likely to increase in response to industrial and agricultural development (Kuylenstierna et al., 2001; Dentener 

et al., 2006). Further emission increases are also occurring in remote areas due to mining activity, such as oil 

sands extraction in Canada (Kelly et al., 2010; Whitfield et al., 2010). 

Sulphur deposition 

Sulphur emissions primarily result from combustion of coal and oil and are typically associated with power 

generation and heavy industry. In 2001, deposition exceedances of 20 kg S ha -1 yr -1 were detected in regions 

of China and Republic of Korea, Western Europe and eastern North America (Vet et al., 2014; Figure 4.10.(a)). 

Deposition in unaffected ecosystems is

The increase in soil pH following the reduction of sulphur emissions shows that the acidification process 

is reversible, although the recovery time is highly variable and dependent on soil properties. Some areas 

with organic soils where deposition has declined are showing either slow or no recovery (Greaver et al., 2012; 

Lawrence et al., 2012; RoTAP, 2012). On agricultural soils, lime can be applied to increase soil pH. However, 

50-80 percent of sulphur deposition on land is on natural land (Dentener et al., 2006). Application of lime to 

naturally acidic forest soils can cause further acidification of deep soil layers by increasing the decomposition in 

surface litter (Lundström et al., 2003). In acid waters, the addition of liming material may favour the formation 

of polymeric Al hydroxides (e.g. Al 13 

OH 27 

+12), which are highly toxic to aquatic species (Monterroso, Alvarez and 

Macías, 1994). 

Wider effects of acidification are starting to be understood through long-term monitoring. Decreased 

organic matter decomposition due to acidification has increased soil carbon storage in tropical forests (Lu 

et al., 2014). In wetland soils, methane (CH 4 

) emissions have also been suppressed. This is because sulphatereducing 

bacteria have a higher affinity for substrate (H 2 

and acetate) than methanogenic microbes (Gauci 

et al., 2004). Conversely, declining sulphur deposition has been associated with increased dissolved organic 

carbon fluxes from organic soils (Monteith et al., 2007) and decreased soil carbon stocks in temperate forest 

soils (Oulehle et al., 2011; Lawrence et al., 2012). 

Nitrogen deposition 

Nitrogen deposition covers a wider geographical area than sulphur deposition. This is because the sources 

are more varied, including extensive agriculture fertilizer and animal waste application, biomass burning, and 

fossil fuel combustion (Figure 4.10c). Regions with deposition in excess of 20 kg N ha -1 yr -1 in 2001 include 

Western Europe, South Asia (Pakistan, India, Bangladesh) and eastern China (Vet et al., 2014). In addition, 

extensive areas with deposits of 4 kg N ha -1 yr -1 or more were found across North, Central and South America 

and parts of Europe and Sub-Saharan Africa. By contrast, ‘natural’ deposition in un-impacted areas is as little 

as 0.5 kg N ha -1 yr -1 (Dentener et al., 2006). While both nitrogen and sulphur emissions related to fossil fuel 

combustion have declined across Europe, agricultural sources of nitrogen in the region are likely to stay 

constant in the near future (EEA, 2014). At the same time, overall global emissions are likely to increase 

(Galloway et al., 2008). Nitrogen deposition in China in the 2000s was similar to peaks in Europe during the 

1980s before Europe embarked on mitigation measures (Liu et al., 2013b). 

Deposition of nitrogen induces a ‘cascade’ of environmental effects, including acidification and 

eutrophication that can have both positive and negative effects on ecosystem services (Galloway et al., 

2003). Soils with low nitrogen content are most sensitive to eutrophication - typically Histosols, Cryosols and 

Podzols located in cold areas in northern countries such as northern Canada, Scandinavia and northern Russia 

(Bouwman et al., 2002; Rodríguez-Lado, Montanarella and Macías, 2007; Figure 4.10d). Excluding agricultural 

areas where nitrogen deposition is beneficial, 11 percent of the world’s natural land experiences nitrogen 

exceedances above 10 kg N ha -1 yr -1 (Dentener et al., 2006). In Europe, eutrophication has and will continue to 

impact a larger area than acidification (Rodríguez-Lado and Macias, 2005; EEA, 2014). 

Nitrogen fertilisation can increase tree growth (Magnani et al., 2007) and cause changes in plant species 

and diversity (Bobbink et al., 2010). This can in turn alter the amount and quality of litter inputs to soils, notably 

the C:N ratio and soil-root interactions (RoTAP, 2012). However, increased global terrestrial carbon sink can be 

largely offset by increased emissions of the greenhouse gases N 2 

O and CH 4 

(Liu and Greaver, 2009). Longterm 

changes caused by nitrogen deposition are uncertain as transport times vary between environmental 

systems. The only way to remove excess nitrogen is to convert it to an unreactive gas (Galloway et al., 2008). 


74

Figure 4.10 Global distribution of (a) atmospheric S deposition, (b) soil sensitivity to acidification, (c) atmospheric N deposition, and 

(d) soil carbon to nitrogen ratio (soils most sensitive to eutrophication have a high C:N ratio; eutrophication is caused by N). Source: 

Vet et al., 2014; Batjes, 2012; FAO, 2007. 

Atmospheric deposition data in (a) and (c) were provided by the World Data Centre for Precipitation 

Chemistry (http://wdcpc.org, 2014) and are also available in Vet et al. (2014). Data show the ensemble-mean 

values from the 21 global chemical transport models used by the Task Force on Hemispheric Transport of Air 

Pollution (HTAP) (Dentener et al., 2006). Total wet and dry deposition values are presented for sulphur, oxidized 

and reduced nitrogen. Soil data in (b) and (d) were produced using the ISRIC-WISE derived soil properties (ver 

1.2) (Batjes, 2012) and the FAO Digital Soil Map of the World. 

Trace element deposition 

Global trace element emissions and deposition are poorly understood in comparison to our understanding 

of emissions of sulphur and nitrogen. Emissions of trace elements are associated with combustion of fossil fuel 

(V, Ni, Hg, Se, Sn), traffic (Pb), insecticides (As), steel manufacture (Mn, Cr), and mining and smelting (As, Cu, 

Zn, Hg) (Mohammed, Kapri and Goel, 2011). In the United Kingdom, trace element deposition is responsible for 

25-85 percent of total trace element inputs to soils (Nicholson et al., 2003). In Europe, the area at risk from Cd, 

Hg and Pb deposition in 2000 was 0.34 percent, 77 percent and 42 percent respectively, although emissions are 

declining (Hettelingh et al., 2006). In China, 43-85 percent of total As, Cr, Hg, Ni and Pb inputs to agricultural 

soils originate from atmospheric deposition (Luo et al., 2009). In bioavailable form these elements have a toxic 

effect on soil organisms and plants, influencing the quality and quantity of plant inputs to soils and the rate of 

decomposition. Significantly, they can also bioaccumulate in the food chain. Activity of trace elements in soils 

will depend on the specific mobility of the element and this will be influenced by pH, Eh and the concentration 

of dissolved organic matter with complexing properties (Blaser et al., 2000). Some trace elements will persist 

for centuries as they are strongly bound to soil particles. However, they can become bioavailable, as observed 

in peatlands following drought-induced acidification, drainage and soil erosion (Tipping et al., 2003; Rothwell 

et al., 2005). 


75

4.4.3 | Knowledge gaps and research needs 

Atmospheric pollution is a global phenomenon impacting large areas of the land surface. Regional and 

global scale assessment relies on the use of simple models to: (i) upscale site-specific soil data, in some 

instances using soil databases collected as long ago as the 1970s; and (ii) estimate where soil sensitivity – the 

‘critical load’ − of a single pollutant is exceeded. There are few locations with long-term soil monitoring data, 

particularly in comparison to the data available on air, rain and surface water quality. Therefore, the actual 

global extent and magnitude of polluted soils are unclear. Essentially, we lack data at adequate scales to check 

the model outputs. A long-term global soil monitoring network is needed. 

While the direct impacts of sulphur, nitrogen and trace elements on inorganic soil chemical processes are 

generally well understood, many uncertainties still exist about pollutant impacts on biogeochemical cycling, 

particularly interactions between organic matter, plants and organisms in natural and semi-natural systems 

(Greaver et al., 2012). Process understanding is dominated by research in Europe and North America (e.g. 

Bobbink et al., 2010). Research is needed in other regions where soil properties and environmental conditions 

differ from the empirically studied areas in Europe and North America. Models need to be developed to 

examine the combined effects of air pollutants and their interactions with climate change and feedbacks on 

greenhouse gas balances and carbon storage (Spranger et al., 2008; RoTAP, 2012). Air quality, biodiversity and 

climate change polices all impact on soils. A more holistic approach to protecting the environment is needed, 

particularly as some climate change policies (e.g. biomass burning, carbon capture and storage) have potential 

to impact air quality and, therefore, soil functions (Reis et al., 2012; RoTAP, 2012; Aherne and Posch, 2013). 

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Global Soil Change 

Drivers, Status and Trends 


André Bationo (Burkina Faso), Srimathie Indraratne (Sri Lanka) 


Sayed Alavi (ITPS/Iran), Abdullah Alshankiti (ITPS/Saudi Arabia), Dominique Arrouays (ITPS/France), Charles 

Bielders (United States), Keith Bristow (Australia), Marta Camps Arbestain (ITPS/New Zealand), Lucrezia 

Caon (Italy), Brent Clothier (New Zealand), Tandra Fraser (United States), Ciro Gardi (Italy), Gerard Govers 

(Belgium), Roland Hiederer (Germany), Jeroen Husing (TSBF-CIAT), Joyce Jefwa (TSBF-CIAT), Shawntine Lai 

(United States/Taiwan, Province of China), Rattan Lal (United States), John P. Lamers (Germany), Dar-Yuan 

Lee (Taiwan, Province of China), Fredah Maina (Kenya), Luca Montanarella (ITPS/EC), Joseph Mung'atu (TSBF- 

CIAT), Freddy Nachtergaele (Belgium), Peter F. Okoth (TSBF-CIAT), Asad Qureshi (ICBA), Shabbir Shahid 

(ICBA), Manuela Ravina da Silva (Sweden), Justin Sheffield (United Kingdom), Tran Tien (Vietnam), Kristof Van 

Oost (Belgium), Boris Vrscaj (Slovenia), Diana Wall (United States), Boaz Waswa (Kenya), Jeewika Weerahewa 

(Sri Lanka), Kazuyuki Yagi (ITPS/Japan), Ted Zobeck (United States). 


Dominique Arrouays (ITPS/France), Richard Bardgett (United Kingdom), Marta Camps Arbestain (ITPS/New 

Zealand), Tandra Fraser (Canada), Ciro Gardi (Italy), Neil McKenzie (ITPS/Australia), Luca Montanarella (ITPS/ 

EC), Dan Pennock (ITPS/Canada) and Diana Wall (United States). 


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5 | Drivers of global soil change 

Drivers in general comprise the factors that bring about socio-economic and environmental changes. They 

operate at various spatial and temporal levels in society. They differ from one region to another, and within and 

between nations. Drivers are diverse in nature and they include: demographics; economic factors; scientific 

and technological innovation; markets and trade; wealth distribution; institutional and socio-political 

frameworks; value systems; and climate and climate change (UNEP, 2007; IAASTD, 2009). Drivers have an 

impact on natural resources including soil services and functions, with impacts on biodiversity, environmental 

health and ultimately human well-being. Globalization has particularly affected these drivers, leading to an 

increase in human mobility with social, economic and environmental implications. Patterns of settlement 

and consumption result in pressures on ecosystem services, including those provided by soils. Rural-urban 

migration and associated livelihood changes contribute to changing patterns of energy use and shifts in diet 

– for example, towards meat – which can intensify pressures on land and soils in producing areas (UNEP, 

2012). In addition, climate change may have significant impacts on soil resources through changes in water 

availability and soil moisture, as well as through sea level rise (IPCC, 2014b). 

5.1 | Population growth and urbanization 

5.1.1 | Population dynamics 

Changing global population trends 

The world population of 7.2 billion in mid-2013 is projected to increase by almost one billion by 2025. By 

2050 it is expected to reach 9.6 billion, and to rise to 10.9 billion by 2100 (UN, 2014). The principal factor in this 

continual rise is the rapid increase in the population of developing countries, in particular in Africa, where the 

population is projected to increase from the current 1.1 billion to reach 2.4 billion by 2050 (Table 5.1.). Many 

countries of Sub-Saharan Africa are still experiencing fast population growth with high fertility rates. Other 

countries with similar trends include India, Indonesia, Pakistan, the Philippines and the United States. By 2030 

Status of the World’s Soil Resources | Main Report Drivers of global soil change 

89

India’s population is expected to surpass China’s, to become the most populous country in the world. Nigeria’s 

population is expected to surpass the United States population in 2045 to become the world’s third most 

populous country. Nigeria’s population is likely to rival that of China by the end of the century (Table 5.2). Over 

the period 2013–2100, eight countries are expected to account for over half of the world’s projected population 

increase: Nigeria, India, the United Republic of Tanzania, the Democratic Republic of Congo, Niger, Uganda, 

Ethiopia and the United States of America. On the other hand, Europe’s population is projected to decline, 

since fertility rates are far below the level for replacement of population in the long run. As fertility decreases 

and life expectancy rises, population ageing is a challenge for Europe (UN, 2014). Other developing countries 

with young populations but lower fertility (e.g. China, Brazil and India) are also likely to face challenges of an 

ageing society by the end of this century (Gerland et al., 2014). 

Region 

Area 

2013 

Percent 

Density 

2050 

% of world 

% Change 

(millions of 

Population 

of world 

(p/km²) 

population. 

pop. 

2013-2050 

km²) 

(millions) 

population 

(projected) 

Asia 31.9 4 298 60.0 135 5 164 54.1 20 

Africa 31.0 1 110 15.5 36 2 393 25.1 115 

Europe 23.0 742 10.4 32 709 7.4 -4 

LAC 20.5 617 8.6 30 782 8.2 27 

North 

America 

21.8 355 5.0 16 446 4.7 26 

Oceania 8.6 38 0.5 4 57 0.6 48 

World 136.8 7 162 100.0 52 9 551 100 33 

Table 5.1 World population by region 

Country 

Population 

Country 

Population 

Country 

Projected 

Country 

Proj. Popul. 

in 1950 

in 2013 

Population 

2100 

in 2050 

China 544 China 1 386 India 1 620 India 1 547 

India 376 India 1 252 China 1 385 China 1 086 

United 

States 

158 

United 

States 

320 Nigeria 440 Nigeria 914 

Russian 

Federation 

103 Indonesia 250 

United 

States 

401 

United 

States 

462 

Japan 82 Brazil 200 Indonesia 321 Indonesia 315 

Indonesia 73 Pakistan 182 Pakistan 271 Tanzania 276 

Germany 70 Nigeria 174 Brazil 231 Pakistan 263 

Brazil 54 Bangladesh 157 Bangladesh 202 

Dem. Rep of 

Congo 

262 

United 

Kingdom 

51 

Russian 

Federation 

143 Ethiopia 188 Ethiopia 243 

Italy 46 Japan 127 Philippines 157 Uganda 205 

Table 5.2 The ten most populous countries 1950, 2013, 2050 and 2100 (population in millions). Source: United Nations, 2014. 


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5.1.2 | Urbanization 

In tandem with the rate of population increase is the rising rate of urbanization. According to the United 

Nations, by 2014 more people were living in urban areas (54 percent) than in rural areas. The urbanization 

trend is expected to continue in all regions, and by 2050, 66 percent of the world’s population is projected to 

be urban. Today some of the most urbanized regions are Northern America (82 percent), Latin America and 

the Caribbean (80 percent) and Europe (73 percent). However, Africa and Asia are now the fastest urbanizing 

regions, with the share of the population urbanized expected to rise from today’s 40 and 48 percent 

respectively to 56 and 64 percent by 2050. Three countries together are expected to account for 37 percent of 

the growth of the world’s urban population between 2014 and 2050: India (adding 404 million), China (adding 

292 million) and Nigeria (adding 212 million). Whereas in the past mega-cities were located in more developed 

regions, today’s large cities are principally found in lower income countries. Since 1990 the number of these 

mega-cities has nearly tripled globally; and by 2030, the world is projected to have 41 global agglomerations, 

housing more than 10 million inhabitants each. In developing countries, the competition between demand for 

agricultural land and the needs of growing cities is a mounting challenge (Jones et al., 2013). 

The rural population globally is now close to 3.4 billion but is expected to decline to 3.2 billion by 2050. 

Africa and Asia are home to nearly 90 percent of the world’s rural population. India has the world’s largest 

rural population (857 million), followed by China (635 million). Rural/urban migration continues to feed urban 

growth, causing environmental changes including effects on land use and soils. Policy and poverty also drive 

the threats to land and soils. Many rural poor live under regimes of weak land policy and insecure tenure 

systems. They often farm marginal lands of low agricultural productivity, typically employing traditional 

farming methods. This may aggravate soil degradation and biodiversity decline, with resulting yield losses 

and food insecurity (Jones et al., 2013; Barbier, 2013). In addition, land grabs may lead to eviction of farming 

families, for whom rural to urban migration may be the only option, so accelerating the pace of urbanization 

(Holdinghausen, 2015). 

5.2 | Education, cultural values and social equity 

Education influences decisions regarding land use and land management. Farmers’ decisions result from 

many factors, including incentives, access to capital and risk management, but also from knowledge and 

level of education, all of which may affect land use and management practices (MA, 2005). Land use and 

management is dependent on the sum total of all decisions taken by individual farmers of different education 

and gender groups in a community (IAASTD, 2009). 

Women play a key role in agriculture. They represent 43 percent of the agricultural labour force world-wide, 

ranging from around 20 percent in Latin America to 50 percent in parts of Africa and Asia (FAO, 2011b). Women 

are responsible for half of the world’s food production, providing between 60 and 80 percent of the food in 

most developing countries (World Bank/FAO/IFAD, 2009). However, evidence shows that women still own 

less land and have smaller landholdings with generally poor soil quality. Improving women’s access to land 

and secure tenure can have direct impacts on farm productivity and in the long run improve household welfare 

(FAO, 2013). FAO’s Gender and Land Rights Database (2010) suggests that less than one quarter of agricultural 

land holdings in developing countries are operated by women. Latin America and the Caribbean have the 

largest mean share of female agricultural land holders, exceeding 25 percent in Chile, Ecuador and Panama. 

In North Africa and West Asia, female landholders represent fewer than five percent of the owners. In sub- 

Saharan Africa the average rate is 15 percent, although there are wide variations within the region, from less 

than five percent in Mali to over 30 percent in Botswana, Cape Verde and Malawi (Figure 5.1). 


91

Percentage of female landholders 

> 29 22.3 - 29 

15.7 - 22.3 9 - 15.7 < 9 No data 

Figure 5.1 Percentage of female landholders around the world. Source: FAO, 2010. 

5.3 | Marketing land 

Today land is used more intensively than ever. The expansion of markets, rising population, and economic 

development and higher incomes have pushed up demand for land for both agriculture and for settlements 

and so driven unprecedented land use change (Section 4.1, this volume). The most dramatic changes have been 

in reduction in forest cover, in expansion and intensification of cropland, and in urbanization (UNEP, 2007). In 

agriculture, production of crops and livestock products for markets is fundamental to the economies of many 

countries. One new segment of market-driven production is biofuels, where incentives have strengthened as 

a result of higher and volatile oil prices and because a number of countries have introduced renewable energy 

promotion policies (Rulli et al., 2013). North America is leading global biofuel production, with 48 percent 

of the global market. The second largest producer of biofuels is Brazil, producing 24 percent of the world’s 

biofuels (OECD/FAO, 2011). The growth of biofuels production is driving an increase in deforestation and other 

land use changes. 

It is widely accepted by economists that when land markets function in an efficient manner, the resulting 

land use patterns provide the highest possible benefits to the society. However, empirical research findings 

reveal that the functioning of land markets in many developing countries is inefficient and the resulting land 

use patterns are sub-optimal (Pinstrup-Anderson and Watson II, 2011). Amongst the causes of inefficiencies 

the following have been cited: lack of well-defined property rights (Allen, 1991; Alston, Libecap and Schneider, 

1995; Besley, 1995); higher bargaining power exercised by different groups of buyers (Sengupta, 1997; Ghebru 

and Holden, 2012); non-existence or under-functioning of insurance markets to absorb risk and uncertainties in 

the natural environment (such as climate change) (Dayton-Johnson, 2006; Auffret, 2003); and environmental 

externalities like soil erosion. 

Land grabbing - large scale land acquisitions - started initially in response to the 2007-2008 increase in food 

prices. Since then the phenomenon has intensified (IMF, 2008). Foreign states and companies and national 

investors, often with the support of the national government, see land as an attractive asset in order to meet 

the demands of food supply and energy. Experience in Africa, Eastern Europe, South America and South and 

Southeast Asia has shown that in an unregulated environment this ‘land grab’ can lead to the displacement 

of local farmers (Rulli et al., 2013). Since fertile land is a limited resource, competition for it may lead to a rise 

in poverty, violence and social unrest in countries with weak regulatory systems or power imbalances (Nolte 

and Ostermeier, 2015). 


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Large areas of arable land have been bought or leased in recent years, mainly in developing countries (Figure 

5.2). According to the Land Matrix Global Observatory database, since the year 2000 over 1 000 land deals 

involving foreign investors have been struck, covering 39 million ha, while another 200 deals cover 16 million 

ha. The main driver of large land-scale acquisitions continues to be agricultural production, with 40 percent 

of deals for food crop production and livestock farming, followed by agrofuels as the second most important 

driver with 190 deals, and forestry projects which have increased by 50 percent (Land Matrix Newsletter, 2014). 

Other acquisitions have been for urban expansion, mining, infrastructure projects and tourism (Nolte and 

Ostermeier, 2015). 

Figure 5.2 Major land deals occurring between countries in 2012. Source: Soil Atlas, 2015/Rulli et al., 2013. 

In addition to these commercial land transactions, policy responses for climate change adaptation and 

mitigation have led to market-based approaches which attach a value to ecosystem services. In this context, 

there is the allocation of land for environmental ends, for example, offsetting emissions in the industrialized 

North by protecting forests in the South. These approaches in practice have sometimes required curtailment 

of customary or community access rights to forest and water. In other cases, these approaches have 

encouraged the shift of smallholder labour from subsistence farming and cash crop production to carbon 

sequestration (UNEP, 2012). A number of projects focusing on soil health and carbon sequestration in Africa 

have aimed to benefit individual farmers and at the same time to mitigate climate change. These pioneering 

projects have faced implementation challenges, including high unit costs, small land sizes, and weak land 

tenure rights. In addition, the incentive framework has been weakened by the small size of the cash payments 

the projects can offer for carbon sequestration, due to the low value of carbon credits and periodic market 

volatility in international voluntary carbon markets. Nevertheless, the non-carbon benefits gained from the 

projects, such as improved agricultural productivity through sustainable land management and soil health 

and strengthening of community solidarity, are important results to be prioritized in future projects (Shames, 

2013). 


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5.4 | Economic growth 

Economic growth and urbanization generate immense benefits to humankind but also contribute 

to unsustainable consumption patterns. They may lead to increased levels of emissions from mining, 

manufacturing, sewage, energy and transport and to the consequent release of persistent pollutants to land, 

air and water (UNEP, 2012). By 2050 the population around the globe is expected to be generally wealthier and 

more urbanized, resulting in an increase and shift in consumption and food demand and consequently in a rise 

in pollution risk. In developing economies, livestock production is already increasing at a rapid rate as a result 

of structural change in diets and consumption. FAO predicts that the total demand for animal products will 

increase at more than double the rate of increase in demand for food of vegetable origin, such as cereals (FAO, 

2011a). This will lead to the expansion of land dedicated to livestock, both pasture and feed production. The 

largest expansion is predicted in the tropics, particularly in South America and Africa where vast areas of tropical 

forests, semi-arid lands, savannah, grassland and wetland ecosystems could be exploited for livestock - with 

potentially devastating environmental results (Laurance, Sayer and Cassman, 2014). In developing countries, 

difficult decisions on the trade-offs between preserving natural ecosystems and economic development will 

be required. In any case, it is likely that agricultural expansion and biofuel production will continue to trigger 

deforestation, and consequently soil degradation, pollution of land and water and increase in greenhouse gas 

emissions (FAO, 2003; Alexandratos and Bruinsma, 2012). 

Despite improvements in income growth in many countries, poverty and access to food remain problematic. 

According to FAO (2014), an estimated 842 million people around the world are currently undernourished. 

The 2007-2008, 2010 and 2012 price hikes in commodity markets evidenced how price shocks can trigger 

prolonged crises leading to food insecurity amongst the most vulnerable (FAO, 2011b). Global agendas such 

as those stated in the Sustainable Development Goals and the Post-2015 Agenda argue that environmental 

stewardship and sustainable management of natural resources provide opportunities to decrease inequality 

while increasing production of goods and services. However, this is a complicated agenda, as links between 

human well-being and natural resources, including soils, are influenced by a host of factors, including 

economic wealth, trade, technology, gender, education etc. Turning these lofty themes into real policies and 

programs to reduce poverty equitably and sustainably remains the key development challenge for the coming 

decades (UNEP, 2007). 

5.5 | War and civil strife 

In the course of history, many conflicts over fertile land have occurred. Until the twentieth century, most of 

these conflicts were local and had relatively little impact on the soils themselves. However, modern warfare 

makes use of non-degradable weapons of destruction and of chemicals that may remain in the affected soils 

for centuries after the conflict. The impacts of war and civil strife on the environment in general, and on soils 

in particular, are both direct physical impacts and indirect socio-economic impacts. 

Direct physical impacts of war on the soil resource include weapons and bombs remaining in the soil, 

the destruction of structures with consequent terrain deformation, heavy military transport that results in 

compaction, and chemical spraying that leads to contamination of both soils and groundwater. Socioeconomic 

impacts of war include local desertification and displacement of large populations of refugees towards safe 

regions, resulting in pressure on the environment and soils in the receiving sites (Owona, 2008). 

Extensive areas in the world are still affected by remnants of past and present war events. Especially affected 

are zones where land mines have been buried (Box 5.1), making these soils unsuitable for any exploitation 

and provisioning of services. There are approximately 110 million mines and other unexploded ordnance (UXO) 

scattered in sixty-four countries on all continents, remnants of wars from the early twentieth century to the 

present (Kobayashi, A., 2013). Africa alone has 37 million landmines in at least 19 countries. Angola is by far the 


94

most affected zone with 15 million landmines and an amputee population of 70 000, the highest rate in the 

world. The problem persists despite campaigns to raise awareness, including the International Campaign to 

Ban Landmines (ICBL), which was awarded the 1997 Nobel Peace Prize. Removal of mines is proceeding, but at 

a glacial pace due to the danger, the cost (US$300 to US$1 000 per mine removed), and lack of international 

agreement on priorities. 

Box 5.1 | Minefields 

Minefields are one of the main constraints to the development of rural areas in Bosnia and Herzegovina 

(BIH), where large tracts (ca. 4 000 km 2 ) of agricultural land and forest areas cannot be used because 

they remain mined after the war that ended two decades ago (ICBL, 2002). BIH is the most mine-affected 

country in Europe with an estimated one million mines (mostly antipersonnel) remaining in the soil, and 

only 60 percent of which are located (Bolton 2003; Mitchell, 2004). This affects about 1.3 million people, 

roughly one third of the population. At current rates of de-mining, it will take several generations before 

rural areas are again safe. 

The displacement of people as a result of wars and conflicts has also created severe environmental and soil 

problems (Box 5.2). 

Box 5.2 | Migration/Refugee Camps 

Acholiland in northern Uganda has suffered from persistent insecurity since the mid -1 980s. The massive 

disruption, dislocation and displacement and suffering of the people in the region are well-known. As a 

way of protecting the local people, the government placed most inhabitants of those districts in camps 

popularly referred to as Internally Displaced Peoples (IDP) camps. As a result, land has been abandoned 

and farming and other socio-economic activities are only possible near the protected camps in a restricted 

radius not exceeding seven kilometres. War creates refugees, leaves government and environmental 

agencies impaired or destroyed, and substitutes short-term survival for longer-term environmental 

considerations. This means that ecosystems continue to suffer even after the fighting has stopped 

(Owona, 2008). The Uganda analysis shows that the creation of 157 IDP camps has significantly affected 

the environment in terms of deforestation (140 km 2 ), soil erosion, habitat destruction and pollution 

(Owona, 2008). 

Often war results in a combination of negative effects on soils. These may include soil compaction, soil 

contamination, soil sealing and enhanced wind erosion and dust fall out (Box 5.3). 

Box 5.3 | Combined effects of war and strife on soils. 

During the 1991 Gulf War in Iraq and Kuwait, there were massive impacts on the environment, resulting 

from heavy vehicle movements, hundreds of oil well fires, numerous oil lakes and spill-outs (Stephens and 

Matson, 1993; El-Baz and Makharita, 1994; El-Gamily, 2007). The desert ecosystem was severely damaged 

by the war: the rate of sand dune movement increased while in addition, new sand sheets and sand dunes 

were formed in several areas where there had been no sheets or dunes previously (Misak et al., 2002; Misak, 

Al-Ajmi and Al-Enezi, 2009). The building of many fortifications exposed huge amounts of fine particles to 

wind erosion Off-site impacts included an increase in the rates of sand transport and dust fallout (Misak, 

Al-Ajmi and Al-Enezi, 2009). The damage remains years afterwards below the surface. 


95

5.6 | Climate change 

The IPCC Fifth Assessment Report reveals that the globally-averaged combined land and ocean surface 

temperature data show a linear trend of global warming due to increases in anthropogenic emissions of 

greenhouse gases of 0.85°C (90 percent uncertainty intervals of 0.65 to 1.06°C). Human influence has been 

detected in warming of the atmosphere and the ocean, in changes in the global water cycle, in reductions 

in snow and ice, in global mean sea level rise, and in changes in some climate extremes. Continued emission 

of greenhouse gases will cause further warming and long-lasting changes in all components of the climate 

system, increasing the likelihood of severe, pervasive and irreversible impacts for people and ecosystems 

(IPCC, 2014a). 

Climate change will have significant impacts on soil resources and food production in both irrigated and 

rainfed agriculture across the globe. Changes in water availability due to changes in the quantity and pattern 

of precipitation will be a critical factor (Turral, Burke and Faurès, 2011). Also, higher temperatures, particularly 

in arid conditions, entail a higher evaporative demand. Where there is sufficient soil moisture, for example in 

irrigated areas, this could lead to soil salinization if land or farm water management, or irrigation scheduling 

or drainage are inadequate. The amount of water stored in the soil is fundamentally important to agriculture 

and is an influence on the rate of actual evaporation, groundwater recharge, and generation of runoff. Soil 

moisture contents directly simulated by global climate models give an indication of possible directions of 

change (IPCC, 2014b). For example, using the HadCM 2 

climate model, Gregory, Mitchell and Brady (1997) show 

that a rise in greenhouse gas concentrations is associated with reduced soil moisture in Northern Hemisphere 

mid-latitude summers. This results from higher winter and spring evaporation caused by higher temperatures 

and reduced snow cover, and from lower rainfall inputs during summer. 

The local effects of climate change on soil moisture, however, will vary not only with the degree of climate 

change but also with soil characteristics (IPCC, 2014b). The water-holding capacity of soil will affect possible 

changes in soil moisture deficits; the lower the capacity, the greater the sensitivity to climate change. Climate 

change also may affect soil characteristics, perhaps through changes in waterlogging or cracking, which in 

turn may affect soil moisture storage properties. Infiltration capacity and water-holding capacity of many 

soils are influenced by the frequency and intensity of freezing. Boix-Fayos et al. (1998), for example, show that 

infiltration and water-holding capacity of soils on limestone are greater with increased frost activity. From 

this, they infer that increased temperatures could lead to increased surface or shallow runoff. Komescu, Erkan 

and Oz (1998) assess the implications of climate change for soil moisture availability in southeast Turkey, 

finding substantial reductions in availability during summer. 

The probable effects on soil characteristics of a gradual eustatic rise in sea-level will vary from place to place 

depending on a number of local and external factors, and interactions between them (Brammer and Brinkman, 

1990). In principle, a rising sea level would tend to erode and move back existing coastlines. In coastal lowlands 

which are insufficiently defended by sediment supply or embankments, tidal flooding by saline water will tend 

to penetrate further inland than at present, extending the area of perennially or seasonally saline soils. 

Climate change such as uncharacteristic droughts or rainfall and flooding have detrimental influences on 

soil microorganisms, changing the natural growing conditions for a region (Gschwendtner, 2014). 

Soil formation is strongly dependent on environmental conditions of both the atmosphere and the 

lithosphere. Soil temperature is an important factor in this physical, chemical and biological process. Soil 

temperature is also an important parameter for plant growth. For example, excessive high temperature is 

harmful to roots and causes lesions of stems, while extreme low temperatures impede intake of nutrients. 

Extreme low and high soil temperatures also influence the soil microbial population and the rate of organic 

matter decomposition. Recent studies have shown that soil temperature is one of the main climate factors 

that influence CO 2 

emission. High soil temperatures accelerate soil respiration and thus increase CO 2 

emission 


96

(Brito et al., 2005). This has implications for the landscape and for land use. On convex slopes and hilltops, 

emission is greater than in foothills, where temperatures are normally lower. Soil surface wetness and canopy 

cover strongly influence the soil’s energy balance and soil temperature, whereas variation in soil porosity and 

soil thermal conductivity have little effect on soil temperature (Luo, Loomis and Hsiao, 1992). The various ways 

to calculate soil temperature and its trends are discussed in Alavipanah (2006) and Alavipanah et al. (2007). 

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6 | Global soil status, 

processes and trends 

6.1 | Global status, processes and trends in soil erosion 

6.1.1 | Processes 

Soil erosion is broadly defined as the accelerated removal of topsoil from the land surface through water, 

wind or tillage. Water erosion on agricultural land occurs mainly when overland flow entrains soil particles 

detached by drop impact or runoff, often leading to clearly defined channels such as rills or gullies. Wind 

erosion occurs when dry, loose, bare soil is subjected to strong winds. Wind erosion is common in semi-arid 

areas where strong winds can easily mobilize soil particles, especially during dry spells. This dynamic physical 

aeolian process includes the detachment of particles from the soil, transport for varying distances depending 

on site, particle and wind characteristics, and subsequent deposition in a new location, causing onsite and 

offsite effects. During wind erosion events, larger particles creep along the ground or saltate (bounce) across 

the surface until they are deposited relatively close to field boundaries (Hagen et al., 2007). Finer particles (< 

80 µm) can travel great distances, with the finest particles entering global circulation (Shao, 2000). Tillage 

erosion is the direct down-slope movement of soil by tillage implements where particles only redistribute 

within a field. 

Status of the World’s Soil Resources | Main Report Global soil status, processes and trends 

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6.1.2 | Status of Soil Erosion 

Over the last decade, the figures published for water erosion range over an order of magnitude of ca. 20 Gt 

(gigaton) yr -1 to over 200 Gt yr -1 . While this huge variation may at first seem to suggest that our estimates of 

global soil erosion are very uncertain, a more detailed analysis shows that estimates exceeding ca. 50 Gt yr -1 are 

not realistic. In most cases, excessively high estimates can be traced back to conceptually flawed approaches 

and/or inappropriate model applications. Considering only those estimates that are not manifestly affected 

by such problems, the most likely range of global soil erosion by water is 20–30 Gt yr -1 , while tillage erosion 

may amount to ca. 5 Gt yr -1 . 

Total erosion rates for wind erosion are highly uncertain. Estimates of the total amount of dust that is 

yearly mobilized on land place an upper limit on dust mobilization by wind erosion on arable land at ca. 2 

Gt yr -1 . However, wind not only mobilizes dust but also coarser soil particles (sand), implying much higher 

total wind erosion rates. A large number of studies have made global estimates of wind erosion and dust 

transport. Approximately 430 million ha of drylands, which comprise 40 percent of the Earth’s surface (Ravi 

et al., 2011), are susceptible to wind erosion (Middleton and Thomas, 1997). In a survey of global estimates of 

present-climate dust emissions, Shao et al. (2011) described 13 studies that estimated global dust emissions in 

a range from 500 to ~ 3320 Tg yr -1 . Ginoux et al. (2012). The studies used global-scale high-resolution satellite 

imagery to study dust sources. They found that natural dust sources do account for about 75 percent of dust 

emissions and the remaining 25 percent of emissions were attributed to anthropogenic sources. The fraction 

of dust sources was highly variable. For example, although North Africa accounted for about 55 percent of the 

global dust emissions, only 8 percent originated from anthropogenic sources. In contrast, anthropogenic dust 

sources contributed 75 percent of the dust emissions in Australia (Ginoux et al., 2012). 

Translating these global estimates into accurate local soil erosion rates is not straightforward as soil 

erosion is highly variable, both in time and in space. However, typical soil erosion rates by water can be defined 

for representative agro-ecological conditions. Hilly croplands under conventional agriculture and orchards 

without additional soil cover in temperate climate zones are subject to erosion rates up to 10-20 tonnes ha -1 

yr -1 , while average rates are often < 10 tonnes ha -1 yr -1 . Values during high-intensity rainfall events may reach 

100 tonnes ha -1 and lead to muddy flooding in downstream areas. Erosion rates on hilly croplands in tropical 

and subtropical areas may reach values up to 50 -1 00 tonnes ha -1 yr -1 . Average rates, however, are lower and 

often 10-20 tonnes ha -1 yr -1 . These high rates are due to the combination of an erosive climate (high intensity 

rainfall) and slope gradients which are generally steeper than those on cultivated land in the temperate 

zones. The incidence of erosion on steep slopes is due not only to specific topographic conditions, but also 

to the combination of a high population pressure with low-intensity agriculture, leading to the cultivation of 

marginal steeplands. 

Rangelands and pasturelands in hilly tropical and sub-tropical areas may suffer erosion rates similar to 

those of tropical croplands. Due to the lack of field boundaries, which often act as barriers for sediment and 

runoff and promote infiltration, these rangelands may also be particularly vulnerable to gully formation. 

This may not affect topsoil so much but may make land inaccessible and hence unusable. Rangelands and 

pasturelands in temperate areas are characterized by erosion rates which are generally much lower and 

are most often below 1 tonnes ha -1 yr -1 . These rangelands are less intensively used and better managed than 

(sub-) tropical rangelands. 

It is possible to identify the areas in the world where soil erosion by water is problematic based on a 

relatively simple modelling approach combining information on soil type, land use, topography and climate 

(Doetterl, Van Oost and Six, 2012; Van Oost et al., 2007). Soil erosion by water is problematic in much of the 

hilly areas that are used as croplands on all continents, even where there have been significant conservation 

efforts as in the Mid-West of United States. Cropland in Europe is characterized by somewhat lower, yet still 

very significant soil erosion rates (Figure 6.1). 


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Prepared by K. Van Oost 

Figure 6.1 Spatial variation of soil erosion by water. High rates (>ca. 20 t ha -1 y -1 ) mainly occur on cropland in tropical areas. 

The map gives an indication of current erosion rates and does not assess the degradation status of the soils. 

The map is derived from Van Oost et al., 2007 using a quantile classification. 

The redistribution of soil within fields due to tillage erosion rates may lead to (very) high erosion rates 

on convexities (knolls) exceeding 30 tonnes ha -1 yr -1 ; and to deposition rates in hollows and at down slope 

field borders exceeding 100 tonnes ha -1 yr -1 . These rates are not directly comparable to wind or water erosion 

rates, as soil eroded by tillage will not leave the field. However, tillage erosion may significantly reduce crop 

productivity on convexities and near upslope field or terrace borders. 

Evidence of past erosion is extensive. This is demonstrated by wind-blown sands of sandstone bedrock, 

extensive loess accumulations of silt-sized aeolian sediments, and other formerly aeolian-affected landscapes. 

Large areas of sand seas, dune fields and other aeolian features and observations of activity provide further 

evidence of past wind erosion (Figure 6.2). 

USDA estimates place wind erosion rates at ca. 2.5 tonnes ha -1 yr -1 on average over all cropland of the United 

States while the average erosion rate for pastureland is ca. 0.1 tonnes ha -1 yr -1 . There are very few quantitative 

assessments of wind erosion rates on arable land outside of the United States. 


102

Figure 6.2 Location of active and fixed aeolian deposits. Source: Thomas and Wiggs, 2008. 

6.1.3 | Soil erosion versus soil formation 

The accelerated loss of topsoil through erosion from agricultural land was recognized as an important 

threat to the world’s soil resource many decades ago. Furthermore, it was feared that soil was, in many areas, 

eroding much faster than that it could be replaced through soil formation processes. More recent studies have 

confirmed that these early observations were not just perceptions. Estimated rates of soil erosion of arable or 

intensively grazed lands have been found to be 100 -1 000 times higher than natural background erosion rates. 

These erosion rates are also much higher than known soil formation rates which are typically well below 1 

tonnes ha -1 yr -1 with median values of ca. 0.15 tonnes ha -1 yr -1 . The large difference between erosion rates under 

conventional agriculture and soil formation rates implies that we are essentially mining the soil and that we 

should consider the resource as non-renewable. 

The imbalance between erosion rates under conventional agriculture and the rate of soil formation implies 

that conventional agriculture on hilly land is not sustainable because the soil resource is mined and will 

ultimately become depleted. This has most likely already happened in many areas around the Mediterranean 

Sea and in tropical mountain regions. So-called soil loss tolerance levels may help to set objectives for shortterm 

action. However, long-term sustainability requires that soil erosion rates on agricultural land are reduced 

to near-zero levels. 

Figure 6.3 Soil relict in the Jadan basin, Ecuador. 

Photo by G. Govers 

In this area overgrazing led to excessive erosion and the soil 

has been completely stripped from most of the landscape 

in less than 200 years, exposing the highly weathered 

bedrock below. The person is standing on a small patch of 

the B-horizon of the original soil that has been preserved. 

Picture credit: Gerard Govers. 


103

6.1.4 | Soil erodibility 

Soil erodibility refers to the degree or intensity of a soil’s state or susceptibility to being eroded (SSSA, 

2008). A critical review of research into the factors controlling susceptibility of soils to wind erosion (‘soil wind 

erodibility’) has been provided by Webb and Strong (2011). Factors controlling soil wind erodibility include 

physical, chemical and biological characteristics of the soil, including texture, aggregation, stability, crusting, 

the amount of loose erodible sediment available, soil water content, roughness due to surface features 

(including tillage marks and vegetation) etc. The factors controlling soil wind erodibility differ somewhat 

among land uses and management approaches. For example, the factors controlling erodibility on rangeland 

differ from the factors controlling erodibility on farmland. In cropped soils at the field scale, disturbance due 

to tillage modifies the soil surface roughness, the amount and distribution of surface cover, soil water content 

aggregation and other properties, all of which affect soil erodibility for short periods of time (Zobeck, 1991; 

Zobeck and Van Pelt, 2011). In arid and semiarid rangeland ecosystems, wind erosion depends on vegetative 

cover and patchiness (Okin et al., 2009) and on surface soil texture and crusting, characteristics that change 

more slowly unless disturbed. Not only do the factors controlling erosion by wind differ among land uses, 

but differences occur in their spatial and temporal variability. Natural and anthropogenic disturbances such 

as grazing, fire and other activities alter the surface and vegetation on rangeland while crop management 

practices often control erodibility of farmland. Webb and Strong (2011) described the dynamics of soil 

erodibility as a continuum that responds to changes in climate variability and disturbance. Factors such as 

rain and crusting on some soils may initially produce low erodibility that will subsequently increase with 

disturbance and drying. The exact timing and variability of changes in erodibility will vary with inherent soil 

physical properties such as soil texture. 

6.1.5 | Soil erosion and agriculture 

Soil erosion has direct, negative effects for global agriculture. Soil erosion by water induces annual fluxes 

of 23-42 Mt (megaton) N and 14.6-26.4 Mt P off agricultural land. These fluxes may be compared to annual 

fertilizer application rates, which are ca. 112 Tg for N and ca. 18 Tg of P. These nutrient losses need to be replaced 

through fertilization at a significant economic cost. Using a United States farm price of ca. US$ 1.45 per kg of N 

and ca. US$ 5.26 per kg of P implies an annual economic cost of US$ 33-60 billion for N and US$ 77 -1 40 billion for 

P 1 . It is therefore clear that compensation for erosion-induced nutrient losses requires a massive investment in 

fertilizer use. In poor regions such as sub-Saharan Africa, the economic resources to achieve compensations 

for nutrient losses do not exist. As a consequence, the removal of nutrients by erosion from agricultural fields 

is much higher than the amount of fertilizer applied. 

The detrimental removal of soil and nutrients from upland fields may be partly offset through the deposition 

of the eroded soil and nutrients in depositional areas. While this is true, such gains should not be exaggerated: 

the deposition of sediments and nutrients in large floodplains is not directly coupled to actual agricultural 

soil erosion, as in most cases sediments are provided by other sources (natural erosion, landslides) and the 

residence time of such sediments in large river systems is several thousands of years. In other words: the 

sediments that are currently being deposited in the Nile valley are not coming from the soils that are currently 

being eroded in Ethiopia. On a smaller scale, the deposition of eroded sediment may indeed locally increase 

local crop productivity, but such effects may be overshadowed by other factors, such as water availability. 

Soil erosion does not induce an important carbon loss from the soil to the atmosphere: erosion mostly 

induces a transfer of carbon from eroding locations to depositional locations. Net losses are limited as the 

carbon lost at eroding locations is partially replaced through dynamic replacement whereas the soil carbon 

that is deposited in colluvial and alluvial settings may be stored there for several centuries. 

1 www.ers.usda.gov/data-products/fertilizer-use-and-price.aspx#26727 


104

High soil erosion rates will also have significant negative effects over longer time spans: the loss of topsoil 

will result in a reduction in the soil’s capacity to provide rooting space and, more importantly, in the capacity 

to store water that can be released to plants. This may reduce soil productivity. However, these changes occur 

relatively slowly: the reduction in soil water holding capacity and and/or root space accommodation results 

in yield declines of ca. 4 percent per 0.1m of soil lost. Except for areas where erosion rates are very high (e.g. 

exceeding 50 tonnes ha -1 yr -1 or ca. 4 mm yr -1 ) this means that effects of erosion on crop productivity will 

be relatively small on the decennial or centennial time scale, provided that nutrient losses due to erosion are 

compensated. Over longer time spans, however, the effect of these losses becomes very significant. 

On the positive side, transported dust affects distant ecosystems, increasing plant productivity by providing 

nutrients not provided by the parent soil, as seen in Hawaii (Chadwick et al., 1999) and the Amazon (Mahowald 

et al., 2008). Transported dust can also provide chemical constituents that affect soil development, as seen in 

the development of terra rossa soils in Bermuda and Spain (Muhs et al., 2010, 2012). 

6.1.6 | Soil erosion and the environment 

The direct negative effects of soil erosion are not limited to agriculture. The sediment produced by erosion 

also pollutes water streams with sediment and nutrients, thereby reducing water quality. In addition, 

sediment contributes to siltation in reservoirs and lakes. However, not all sediments trapped in reservoirs 

originate from agricultural land. Other processes such as bank erosion, landslides and natural surface erosion 

which contribute to reservoir sedimentation are also very important and are often dominant at large scales. 

Wind erosion and dust transport have been studied for many years. For example, in 1646, Wendelin first 

described purple rain in Brussels that we now recognize as coloured dust transported to Europe from Africa 

(Wendelin, 1646 as cited in Stout, Warren and Gill, 2009). Charles Darwin studied dust that fell on the HMS 

Beagle in the 1830s and 1840s (Darwin, 1845, 1846) and the dust collected was found to contain viable microbes 

even today (Gorbushina et al., 2007). 

Wind erosion can originate from natural landscapes and from landscapes affected by anthropogenic 

(human-related) activities (Figure 6.4). Aeolian processes impact soil development, mineralogy, soil physical 

and biogeochemical properties, and redistribution of soil nutrients, organic materials, and sequestered 

contaminants. Wind erosion also affects landscape evolution, plant productivity, human and animal health 

(Ravi et al., 2011), atmospheric properties including effects on solar radiation and cloud attributes (Shao et al., 

2011), air quality, and other factors (Field et al., 2010; Ravi et al., 2011). The effects of wind erosion occur at the 

field, landscape, regional, and global scales. 


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Photo by T.M. Zobeck 

Figure 6.4 Dust storm near Meadow, Texas, USA 

At the field and landscape scales, wind erosion winnows the finer and more chemically active portion 

of the soil which carries biogeochemicals, including plant nutrients, soil carbon and microbial products. In 

some cases, wind erosion processes modify the surface properties by causing increases in sand content while 

reducing the soil water holding capacity and plant productivity (Zobeck and Van Pelt, 2011). Although some of 

this eroded sediment is deposited relatively close to field boundaries, often much of it enters into suspended 

mode and may be transported high in the atmosphere to travel great distances. This long-range transport of 

dust produces effects at the global and regional scales Atmospheric dust produced by wind erosion profoundly 

influences the energy balance of the Earth system by carrying organic material, iron, phosphorus and other 

nutrients to the oceans, affecting ocean productivity and subsequent ocean-atmosphere CO 2 

exchange (Shao 

et al., 2011). 

6.1.7 | Effects of hydrology and water 

Wind erodibility and subsequent erosion and dust emissions are affected by hydrology and water in several 

ways. Remote sensing studies of dust sources by Prospero et al. (2002) showed that many major dust sources 

originate from deep alluvial deposits formed by intermittent flooding during the Quaternary and Holocene. 

These sources, now in drylands, originated when water was more plentiful and produced an ample supply of 

wind-erodible sediment (Ginoux et al., 2012). In many areas, particularly in areas with more limited erodible 

sediment supply, dust emissions increase after new inundations of ephemeral water supplies provide additional 

erodible sediment. However, many fluvial-related dust sources have also developed from the exposure, due to 

anthropogenic factors, of the bottoms of former lakes such as at Owens Lake in the United States (Reheis, 

1997) and the Aral Sea Basin in Uzbekistan (Singer et al., 2003). In these cases, usually water has been extracted 

from the lake for irrigation or human consumptive needs. This issue will be accentuated as increasing demand 

for water in dryland regions is met from reservoirs. 


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Near-surface soil water content has long been recognized to have a significant effect on the threshold wind 

velocity needed for wind erosion (Akiba, 1933; Chepil, 1956). Soil water acts to bind particles together to resist 

the shearing force of wind on the particles. In addition, soil water affects vegetative growth, which also affects 

wind erosion. Research has shown a time-dependent change in the controlling factors for sediment emission 

and transport from soil water to wind speed (Wiggs, Baird and Atherton, 2004). The change of controlling 

factors was found to be very sensitive to the prevailing water conditions and, for the sandy soil tested, took 

place in a very short period of time. They found the soil water content where wind erosion commenced was 

between 4 and 6 percent (Wiggs, Baird and Atherton, 2004). However, the effect of soil water on wind erosion 

of dry soils is also sensitive to changes in air relative humidity (Ravi et al., 2006). Recent work on atmospheric 

dust concentrations have confirmed this sensitivity, finding that dust concentration increased with relative 

humidity, reaching a maximum around 25 percent and thereafter decreased with relative humidity (Csavina et 

al., 2014). Climate-induced changes in hydrology and water may produce profound changes in wind erosion 

and dust emissions as the soil erodibility is altered. 

6.1.8 | Vegetation effects 

The effect of vegetation on wind erosion is complex. In native conditions, the wind influences patterns 

of vegetation and soils and these patterns, in turn, affect wind erosion at patch to landscape scales (Okin, 

Gillette and Herrick, 2006; Okin et al., 2009; Munson, Belnap and Okin, 2011). In agricultural systems, the 

vegetation is manipulated by managers and its effects vary spatially and temporally from non-managed 

systems. The protective effects of vegetation are well known. A wide variety of methods and models has been 

devised to describe the protective effects of vegetation. In general, as vegetation height and cover increase, 

wind erosion of erodible land decreases. Vegetation affects wind erodibility by: (1) acting to extract momentum 

from the wind and thereby reducing the wind energy applied to the soil surface; (2) directly sheltering the soil 

surface from the wind by covering part of the surface and reducing the leeside wind velocity; and (3) trapping 

windborne particles, so reducing the horizontal and vertical flux of sediment (Okin, Gillette and Herrick, 

2006). Trapping of sediment leads to redistribution of nutrients and modifies surface soil properties such as 

water infiltration rate and soil bulk density. 

Vegetation cover affects nutrient removal, which in turn affects plant productivity. A study of the effects 

of grass cover on wind erosion in a desert ecosystem found increased wind erosion removed 25 percent of the 

total soil organic carbon and nitrogen from the top 5 cm of soil after only three windy seasons (Li et al., 2007). 

Studies of agricultural crops on severely eroded cropland found 40 percent reductions in cotton and kenaf 

yields and 58 percent reduction in grain yield in sorghum (Zobeck and Bilbro, 2001). The eroded areas in this 

study had statistically significantly less phosphorus than the adjacent non-eroded areas. Climatic changes 

that reduce the cover of vegetation in drylands will increase wind erosion and dust emissions, and likely result 

in increased soil degradation and reduced plant productivity. 

6.1.9 | Alteration of nutrient and dust cycling 

Recognition of a dust cycle, along with other important cycles such as the energy, carbon and water cycles, 

has become an emerging core theme in Earth system science (Shao et al., 2011). Dust cycles are dependent 

upon the soil and climate systems within which they operate. The dust cycle is a product, in part, of the soil 

system. As dust is transported globally, it interacts with other cycles by participating in a range of physical 

and biogeochemical processes. The dust carries important nutrients to otherwise sterile soils and so may 

improve productivity (Chadwick et al., 1999; Mahowald et al., 2008). Dust may also transport soil parent 

material (Reynolds et al., 2006), trace metals (Van Pelt and Zobeck, 2007), soil biota (Gardner et al., 2012) and 

toxic anthropogenics (Larney et al., 1999) among ecosystems. Although the fact is not widely recognized, the 

global dust cycle is intimately tied to the global carbon cycle (Chappell et al., 2013). Wind and water erosion 

both redistribute soil organic carbon within terrestrial, atmospheric and aquatic ecosystems. This carbon is 

selectively removed from the soil. This was recently demonstrated in an Australian study where the soil organic 

carbon in dust was from 1.7 to over seven times that of the source soil (Webb et al., 2012). Changing climate will 

alter these cycles, producing complex and uncertain environmental effects. 


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6.1.10 | Trends in soil erosion 

While rates of soil erosion are still very high on extensive areas of cropland and rangeland, erosion rates have 

been significantly reduced in several areas of the world in recent decades. The best documented example is the 

reduction of erosion rates on cropland in the United States. Average water erosion rates on cropland were 

reduced from 10.8 to 7.4 tonnes ha -1 yr -1 between 1982 and 2007, while wind erosion rates reduced from 8.9 to 

6.2 tonnes ha -1 yr -1 over the same time span. Another example is the reduction of soil erosion in Brazil through 

the application of no-tillage from ca. 1980 onwards. This is estimated to have led to a reduction of erosion 

rates by 70-90 percent over large parts of Brazilian cropland. Studies have shown that erosion rates can be 

greatly reduced in nearly every situation through the application of appropriate management techniques and 

structural measures such as terrace and waterway construction (see, for example, Pansak et al., 2008). 

However, in many areas of the world, adoption of soil conservation measures is slow. While the reasons for 

this are diverse, a key point is that the adoption of soil conservation measures is generally not directly beneficial 

to farmers. This is as true in intensive mechanized systems in the West as it is for smallholder farming in the 

developing world. This is not surprising: the implementation of conservation measures does not, as such, 

directly increase yields or efficiencies while the detrimental effects of erosion on the soil capital only become 

visible over time scales that range from decades to centuries. Hence, farmers do not have a direct incentive to 

adopt soil conservation measures. 

In some cases, this problem may be overcome through financial incentives or by regulation. It is clear, 

however, that this is not always possible. We need, therefore, to rethink our vision on soil conservation. 

Essentially, further adoption of soil conservation measures will not in the first place depend on refinement 

or optimization of technologies. Technology already exists to reduce erosion to acceptable levels under 

most circumstances. What is critically important is to work out how to incorporate soil conservation in an 

agricultural system that, as a whole, increases the net returns of farmers. In developing approaches that build 

in incentives to soil conservation, it is vital to account for local conditions, including the extent to which local 

markets can provide incentives to sustainable agriculture. 

The potential for agricultural intensification is key here. In many areas around the world, crop yields are 

low and more land is cultivated than is strictly necessary. As a result, large tracts of steep, marginal land are 

at present used for agriculture without the implementation of proper soil conservation technology, with the 

result that these areas are subject to high erosion rates. Intensification of production on higher potential 

land is an option. This not only reduces extension into marginal, highly erodible areas but may also benefit 

biodiversity and overall carbon storage at the landscape scale. 

Erosion can also be checked by forestation. In many areas there is now a net gain of forest area. This 

reforestation, which is largely of marginal land, is related to four main factors: agricultural intensification; 

diminishing need for firewood; an increase in exchange and trade making it possible to grow products in the 

most suitable areas; and an increased public awareness of the problems caused by deforestation. Development 

of conservation policies should consider these tendencies and stimulate them wherever possible. 

6.1.11 | Conclusions 

Soil erosion has been recognized as a main problem threatening the sustainability of agriculture for a long 

time and the magnitude of the problem can now be correctly quantified. The technology to reduce erosion 

now exists and, over the last decades, significant efforts have been to reduce erosion rates. These efforts have 

been partially successful. However, erosion rates are still high on much of the agricultural land of the globe, 

and this is related to the lack of economic incentives for today’s farmers to conserve the soil resource for future 

generations. Tackling this problem requires the soil erosion problem to be reframed. Solutions need to be 

embedded in policies and programs that support the development of more sustainable agricultural systems. 


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6.2 | Global soil organic carbon status and trends 

6.2.1 | Introduction 

An evaluation of the various functions of carbon (C) stored in the soil and its role in the global C-cycle 

require knowledge of the amount and geographic distribution of C stored in the soil. The functions of soil C are 

determined by the chemical and physical properties of the components that contain C. Chemical properties of 

soil C determine properties such as soil pH, nutrient storage and availability and regulating functions affecting 

the water cycle. Soil physical properties related to functions of C are: soil structure, particle agglomeration, 

and stability. These properties in turn influence water infiltration rates and resistance to water and wind 

erosion. 

Soil C is separated into: (i) inorganic chemical substances (soil inorganic carbon: SIC), mainly carbonates; 

and (ii) C as part of organic compounds (soil organic carbon: SOC). The amount of SIC in the first one meter of 

soil was estimated at 695 - 748 Pg carbonate C (Batjes, 1996). The C stored as SOC is about twice the C stored as 

SIC (1 502 Pg C; Jobbágy and Jackson, 2000). Carbonates are less susceptible to react to anthropogenic changes 

to the environment than are organic compounds. In addition, the amount and type of organic C compounds 

are interdependent with environmental conditions, such as land use and management practices. These two 

characteristics have led to the definition of the persistence of SOC as an ecosystem property (Schmidt et al., 

2011). Thus, assessments of soil C stocks and their spatial distribution often concentrate on SOC alone. 

Although SOC mainly originates from plant material there is no simple correlation between the amount of 

C stored in the above-ground plant material and the SOC stocks (Amundson, 2001; Smith, 2012). In fact, the 

processes involved in decomposing organic material and their mineralization are complex and details are not 

yet fully understood (Schmidt et al., 2011). However, the effects on SOC of anthropogenic activities of land use 

change and management practices are known. Given the large amount of C stored in soils and the possibility 

of influencing this amount through land management to act as a source or sink for atmospheric C, strategies 

for maintaining SOC have been devised. These strategies follow two main approaches: (i) seeking to enhance 

existing SOC by increasing the amount of biomass of the terrestrial biosphere; (ii) seeking to decrease the loss 

of SOC by reducing the respiration rate (Smith et al., 2008). To provide a quantitative appraisal of the possible 

gains or losses in SOC from measures taken either to increase the input of organic material or decrease losses 

of soil organic matter (SOM), an assessment of the current situation is needed. 

Studies on historic developments in SOC stocks concentrate on the effect of changes in land use. These 

changes mainly concern the transformation of land from a natural state to an agricultural ecosystem, which 

in fact now covers more than one third of the global terrestrial area. For the conversion of forest to cropland, 

losses in SOC stocks of 25-30 percent were observed for temperate regions, with higher losses recorded for the 

tropics. Estimates of future trends in SOC stocks concentrate on the effect of changing climatic conditions 

on rates of organic matter accumulation and decomposition. As options for changes in land use are relatively 

limited, approaches to mitigation of climate change effects have focussed largely on management practices. 

6.2.2 | Estimates of global soil organic carbon stocks 

It is important to know past, current and likely future SOC stocks because of their importance to climate 

change and food security. When assessing the amount of C stored in the soil, studies often concentrate on C 

contained in dead and decomposed organic material or in organic matter located within the soil profile to a 

given depth and for a specific area. The mass of C stored in the SOM is also termed SOC stocks. SOC stocks 

are computed as a function of organic C-content, bulk density, depth and the amount of soil remaining after 

removing the volume taken up by coarse fragments in a unit of volume. Any of these factors can introduce 

uncertainties to the estimates. 


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Global estimates of SOC stocks have been published for many decades. One of the earliest estimates was 

published in 1951 (Rubey, 1951), indicating a global SOC stock of 710 Pg C. This estimate remained current for 

25 years until the FAO-UNESCO soil map became available. Analysis of the map data led to a much higher 

estimate of 3 000 Pg C in the soil (Bohn, 1976). Several studies of global SOC stock followed with varying 

estimates (Bohn, 1982: 2 200 Pg organic C; Post et al., 1982: 1 395 Pg organic C) An estimate of 1 576 Pg of SOC 

to 1 m depth was put forward by Eswaran, Van Den Berg and Reich, 1993. Global SOC stocks to 1 m depth of 

1 462 – 1 548 Pg of SOC were estimated by Batjes, 1996. An estimate of 1 502 Pg organic C for the first 1 m of soil 

is often used (Jobbágy and Jackson, 2000; Batjes, 2002). The estimate of 1 500 Pg of SOC for the top 1 m of 

soil was adopted by the IPCC (IPCC, 2000). Current global estimates, derived from the Harmonized World Soil 

Database (HWSD; FAO/IIASA/ISRIC/ISS-CAS/JRC, 2009), suggest that approximately 1 417 Pg of SOC are stored 

in the first meter of soil and about 716 Pg organic C in the top 30 cm (Hiederer and Köchy, 2011). 

Fewer estimates of global SOC stock estimates are available for a depth below 1 m. Global SOC stocks to a 

depth of 3 m are estimated at 2 344 Pg of SOC (Jobbágy and Jackson, 2000) or 3 000 Pg of SOC (Jansson et al., 

2010). Recent estimates of SOC in Cryosols may further increase the estimates of global SOC stocks (Tarnocai 

et al., 2009). 

In a comparison of 27 studies on global SOC stock published between 1951 and 2011, the estimates published 

were found to range from 504 to 3 000 Pg of SOC (Scharlemann et al., 2014). The median of all published 

estimates is 1 460 Pg of SOC to 1 m depth. Spatially explicit estimates were found to span over 1 965 Pg of 

SOC. Large uncertainties over SOC stocks concern Histosols since soil data are often limited to a depth of 1 m 

(Eswaran, Van Den Berg and Reich, 1993). Particularly affected are the soils of the Arctic (Tarnocai et al., 2009) 

and peatlands in South Asia (Couwenberg, Dommain and Joosten, 2010). The range in the estimates of global 

SOC stocks correspond to or exceed the amount of C held in the atmosphere, which was estimated at 720 Pg 

C (Falkowski et al., 2000) and at 820 Pg C for present conditions (Mackey et al., 2013). 

With respect to the uncertainty in the estimates of global SOC stocks, various approximations are observed. 

For an estimated SOC stock of 1 395 Pg of SOC Post et al. (1982) assume a standard deviation of ± 200 Pg 

organic C, provided that the SOC density data are the only source of uncertainty. For the estimate of 1 502 Pg 

organic C to 1 m depth, Jobbágy and Jackson (2000) suggested an error of the mean of ± 320 Pg C at 1 standard 

deviation, provided that the SOC content data are the only source of uncertainty. The different assumptions 

on the causes of uncertainty between the studies (SOC density or content) are quite significant. Based on the 

HWSD, Todd-Brown et al. (2013) provide an interval of estimated global SOC stock of 890 to 1 660 Pg of SOC 

to a depth of 1 m with a 95 percent confidence level. This range corresponds to approximately 385 Pg SOC at 2 

standard deviations from the mean. 

With a small number of large-scale data sets available, the variations in SOC stock estimates may be 

attributed to the analysis method applied as much as to the data used. It also implies that various global SOC 

stock estimates are not independent and that the variability in the estimates could not necessarily be reduced 

by an increase in the number in such estimates. 

One problem common to all large databases is that the properties were assessed decades ago and stretched 

over long periods. For example, the DSMW or the ESDB, of which components are included in the HWSD, 

originated from data collected during the 1950s and 1960s. With the dependence of SOC on climatic conditions 

and anthropogenic activities, SOC stocks established decades apart are likely to represent significantly 

different levels, notably in areas where changes in land use or management occurred, such as conversion of 

natural grassland and forest to agricultural land or urban areas. In extreme cases draining peatlands can lead 

to a loss of organic material to the degree that the soil no longer qualifies as peat because the organic C content 

decreases below 12 percent content and the thickness of the remaining organic layer is less than 40 cm (FAO/ 

ISRIC/ISSS, 1998). An example of this change is given by the agricultural areas in north-eastern Netherlands, 


110

where the drainage in the 1960’s of areas previously classified as peatland caused the SOC content to fall to 7.5 

percent (Panagos et al., 2013). Without further adjustments of SOC stock estimates to take account of local 

changes in the factors that influence SOC, no clear timestamp can be attached to the global estimates. This 

lack of a clear timestamp of SOC stocks is of consequence when estimating temporal changes in SOC stocks. 

Estimates of changes in SOC stock therefore concentrate on modelling variations in SOC from changes in land 

use and cover. 

6.2.3 | Spatial distribution of SOC 

Different methods of combining point data from soil profiles with soil spatial layers and ancillary ecological 

data can be applied to derive spatial estimates of SOC stocks (Kern, 1994). SOC density and stock estimates 

from soil profile data were combined with spatial data of major ecosystems by Post et al. (1982). The total SOC 

stocks for all life zones to a depth of 1 m was 1 395 Pg of SOC. A combination of soil profile data with ancillary 

information on climate, vegetation and land use was used by Jobbágy and Jackson (2000) to estimate SOC 

stocks in 11 biomes. The estimates for the biomes were further divided into increments of 1 m soil depth and 

of 20 cm for the first meter. The distribution of SOC stocks by ecological regions has also been presented, for 

example by Amundson (2001), who used life zones as the study unit. Eglin et al. (2010) used the SOC stock 

estimates to a depth of 3 m from Jobbágy and Jackson (2000) and modified SOC stocks in permafrost areas 

(Tarnocai et al., 2009). These SOC stock estimates were combined with estimates provided by the IPCC (IPCC, 

2000) of C in vegetation to derive estimates of C in soil and biomass for 10 biomes, with an explicit class for 

peatlands. A step towards adding a temporal dimension to spatial SOC stock estimates, assessing historical 

and future trends, was made possible by the availability of SOC models. Combining the models with historic 

land use and climate data has allowed estimation of SOC stocks with a timestamp and with regional variations 

(Eglin et al., 2010; Schmidt et al., 2011). 

Carré et al. (2010) produced estimates of SOC stocks and density using climate data, IPCC methodology and 

the Harmonized World Soil Database. The results by IPCC Climate Region are presented in Table 6.1. 


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IPCC Climate Region IPCC HWSDa 

Tier 1 Topsoil Subsoil Soil Density 

0-30 cm 0-≤30 cm 30-≤100 cm 0-≤100 cm 0-≤30 cm 

Pg C Pg C Pg C Pg C* Mg C ha -1 

Tropical Wet 52.4 62.6 65.4 128.0 66.5 

Tropical Moist 94.5 78.6 72.3 150.9 45.0 

Tropical Dry 99.9 67.3 69.0 136.2 22.0 

Tropical Montane 49.8 29.6 26.5 56.1 40.3 

Warm Temperate Moist 41.7 33.3 29.7 63.0 60.2 

Warm Temperate Dry 42.9 38.9 39.6 78.5 30.8 

Cool Temperate Moist 110.6 104.1 106.2 210.3 88.2 

Cool Temperate Dry 56.9 52.2 50.0 102.2 42.7 

Boreal Moist 137.3 162.0 194.7 356.7 117.6 

Boreal Dry 30.3 32.0 37.0 68.1 84.0 

Polar Moist 26.8 30.6 21.7 52.4 40.4 

Polar Dry 7.2 8.0 4.3 12.3 40.5 

Total 750.3** 699.3 716.4 1415.7 52.1 

Table 6.1 Distribution of Soil Organic Carbon Stocks and Density by IPCC Climate Region 

* Differences in topsoil and subsoil sum are due to data rounding 

** Total includes 1.4 Pg C in undefined climate regions 

The table shows that according to the processed data from HWSD, most SOC (356.7 Pg C) is stored in the 

‘Boreal Moist’ climatic region. The second largest stock is found in the ‘Cool Temperate Moist’ region (210.3 Pg 

C). With 117.6 Mg C ha -1 and 88.2 Mg C ha -1 , these climate regions also have the highest SOC densities. These 

figures compare poorly with those presented by Post et al. (1982). A major source for the deviation is the 

difference in the definition of the life zones as compared to the climatic regions, which lead to the delineation 

of different areas. 

Using the IPCC Tier 1 default values for organic C in mineral soils and retaining the stocks for organic soil 

gives global organic C stock in the upper 30 cm of soil of 750.3 Pg C. This estimate is 51 Pg C (7.3 percent) higher 

than the estimates derived from the HWSDa topsoil layer. 

When comparing the two spatial SOC stock estimates by IPCC climate region, the stocks within each region 

are broadly similar. A notable difference is for soils in the ‘Tropical Dry’ climate region. The IPCC Tier 1 SOC map 

gives 99.9 Pg C for this zone, compared to 67.3 Pg C found in the HWSDa. 

In the interpretation of the figure for SOC stocks of the IPCC Tier 1 Soil Organic Carbon layer, it should be 

considered that organic soils were only added to the mineral soil layer in places where this soil type is not 

found in association with mineral soils. Using all organic soil data is likely to increase the global SOC stocks. 

However, the Tier 1 default values are calculated over a constant depth of 30cm, although some soils are 

shallower, which in turn would reduce the stocks. 


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6.2.4 | Spatial distribution of carbon in biomass 

A global map of C stored in biomass following the IPCC Tier 1 approach was produced by the European 

Commission Joint Research Centre (Carré at al., 2010; EU, 2004; Hiederer et al., 2010). The C stocks are 

determined for above- and below-ground biomass and include dead organic matter for the relevant vegetation 

types. The default factors largely follow the IPCC specification, with specific attention given to agricultural 

areas. The underlying vegetation data are based on the GlobCover V 2.2 (ESA, 2011). Because the GlobCover data 

limits cropland to areas below 57º N in Europe the data were merged with the M 3-Cropland (Ramankutty et 

al., 2008) and Crops (Monfreda, Ramankutty and Foley, 2008). In a comparison of the geographic distribution 

of IPCC vegetation classes between the GlobCover and the M 3 Cropland and Pasture data, some notable 

differences were identified (Hiederer et al., 2010). Some of the differences were attributed to the dissimilar 

definition of the vegetation classes in the data sets, although others, such as the separation of shrub land 

from open forest or confusion between cropland and pastures, seem to be the result of the classification 

algorithm used or of sensor characteristics. 

The global biomass map thus generated by the Joint Research Centre (JRC) estimates the storage of C in 

the above-ground and below-ground vegetation and dead organic matter to be 456 Pg C. The JRC estimates 

are thus 44 Pg C (8.8 percent) lower than those of the ‘New IPCC Tier -1 Global Biomass Carbon Map for the 

Year 2000’ (Ruesch and Gibbs, 2008). The difference is not evenly distributed between geographic regions. A 

comparison of carbon in C by climatic region is given in Figure 6.5. 

Figure 6.5 Distribution of carbon in biomass between ORNL-CDIAC Biomass and JRC Carbon Biomass Map 

The graph shows that the ORNL-CDIAC Biomass and the JRC Carbon Biomass map are mostly comparable, 

but the JRC map places relatively more C in the biomass in ‘Cool Temperate Moist’ (11.4 percent of the total C 

stock in biomass; 51.8 Pg C) and ‘Warm Temperate Moist’ (8.7 percent of the total C stock in biomass; 39.9 Pg 

C) climate regions at the expense of other regions. By contrast, the ORNL-CDIAC Biomass map locates only 5.7 

percent of the total C stock in biomass (28.4 Pg C) in the ‘Cool Temperate Moist’ and 5.7 percent of the total C 

stock in biomass (28.7 Pg C) in the ‘Warm Temperate Moist’ climate region. 

For the total terrestrial pool of organic C, biomass is the more important pool only in the climate regions 

‘Tropical Wet’ and ‘Tropical Moist’. For all other climatic regions, the soil stores more organic C than the biomass 

(Scharleman et al., 2014). 


113

6.2.5 | Distribution of terrestrial carbon pool by vegetation class 

Areas where SOC or biomass C dominate could be identified by computing the difference between the two 

layers. The resulting layer is presented in Figure 6.6. 

Figure 6.6 Prevalence of carbon in the topsoil or biomass 

The figure shows that, as a general propensity, soil dominates the terrestrial C pool in cooler climates while 

vegetation forms the dominant pool of terrestrial C in tropical regions. 

In an attempt to provide C stock estimates for broad land use activities, global GLC 2000 data layers 

were used. The GLC 2000 categories were re-classified according to the assignments for these classes given 

by Ruesch and Gibbs (2008). A difference in the assignment was applied to GLC 2000 classes 16 (Cultivated 

and Managed Areas) and 23 (Irrigated Agriculture). In the broad classification these areas were grouped with 

other areas mainly associated with an absence of soil or biomass (bare areas, glaciers, etc.). For the analysis 


114

Vegetation Classes Topsoil Subsoil Soil Biomass 

Terrestrial 

C Stock 

Pg C Pg C Pg C percent Pg C percent Pg C percent 

Broadleaf Forest 124.7 112.4 237.1 16.8 272.2 54.4 509.4 26.6 

Evergreen Forest 126.8 139.7 266.4 18.8 46.4 9.3 312.9 16.3 

Mixed Forest 40.5 47.8 88.3 6.2 21.8 4.4 110.1 5.7 

Burnt Forest and 

Natural Forest 

Mosaic 

Forest/Cropland 

Mosaic 

27.4 36.2 63.6 4.5 10.9 2.2 74.5 3.9 

23.2 23.4 46.6 3.3 28.0 5.6 74.6 3.9 

Forest 342.6 359.5 702.0 49.6 379.3 75.9 1081.5 56.4 

Shrub Cover 89.2 102.4 191.6 13.5 51.8 10.4 243.4 12.7 

Grasslands 60.5 52.1 112.6 8.0 18.0 3.6 130.5 6.8 

Sparse Grassland 

and Grassland 

Mosaic 

69.0 65.5 134.5 9.5 12.7 2.5 147.2 7.7 

Grassland 218.7 220.0 438.7 31.0 82.5 16.5 521.1 27.2 

Agriculture and 

managed areas 

80.8 79.4 160.2 11.3 26.7 5.3 186.9 9.8 

Other Classes 57.3 57.4 114.7 8.1 11.4 2.3 126.2 6.6 

Figure 6.7 Proportion of carbon in broad vegetation classes for soil and biomass carbon pool 


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of the distribution of organic C, a separate class of ‘Agriculture and Managed Areas’ was created by merging 

the GLC 2000 classes 16 and 23. For each of the broad vegetation classes the organic C stock was extracted by 

pool. The results are presented in Table 6.2. 

The single largest stock for terrestrial C is attributed to areas with broadleaf forest (509.4 Pg C). This forest 

type contains approximately one quarter of all terrestrial organic C in either the soil or the biomass. 

The proportions of the C stored in the soil and biomass stocks by broad vegetation class is graphically 

presented in Figure 6.7. 

For the biomass C stock alone, broad forests account for over 50 percent of the C in that pool, but only 16.8 

percent of the organic C is stored in the soils under this vegetation type. With the exception of the ‘Forest/ 

Cropland Mosaic’, in all other vegetation classes the soil stores more C than the biomass. 

6.2.6 | Historic trends in soil carbon stocks 

The SOC stocks are more susceptible to anthropogenic activities and natural factors than are SIC stocks. 

Conversion of natural to agro-ecosystems in the past has led to decline in the SOC stock of the surface layers 

and also in SOC in the total profile for most soils. The magnitude of the historic loss, however, differs among 

soils and climates. The magnitude and rate of loss are higher for soils within the tropics than for those of 

temperate climates. Losses are also higher for coarse-textured than for heavy-textured soils, higher for 

soils containing higher SOC stocks, and higher for soils under subsistence or ‘extractive’ farming than for 

those farmed with more science-based agricultural practices. Depletion is also exacerbated by drainage of 

wetlands, by ploughing, and by biomass burning or removal. Some soils in the tropics can lose 50 percent of 

their previous pool within five years following deforestation and conversion to agricultural land use. The rate 

and magnitude of SOC loss are exacerbated in soils vulnerable to accelerated erosion, salinization, nutrient 

depletion or imbalance, structural decline and compaction, acidification, elemental toxicity, pollution and 

contamination. 

Estimates of the magnitude of historic SOC loss vary widely. The historic loss has been estimated at 40 Pg by 

Houghton (1995), 55 Pg by IPCC (1995) and Schimel (1995), 150 Pg by Bohr (1978), 500 Pg by Wallace (1994) and 

537 Pg by Buringh (1984). The average of these estimates is 223 Pg C year -1 . Lal (1999) estimated the magnitude 

of SOC loss since 1850 at 47 to 104 Pg for different biomes (Table 6.3); 66 to 90 Pg for major soils (Table 6.4); 

Biome 

Change in Area 

10 6 ha 

Historic SOC Loss 

Pg C 

Forests 1300 23 - 53 

Woodlands 180 3 - 7 

Shrublands 140 1 - 4 

Grasslands 660 20 - 40 

Total 47 - 104 

Table 6.3 Estimate of the historic SOC depletion from principal biomes. Source: Lal, 1999. 


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Soil Order 

Historic Area 

10 6 ha 

Present SOC 

Pool Pg C 


Pg C 

Alfisols 1330 91 15 - 18 

Andisols 110 30 5 - 7 

Aridsols 1560 54 0.2 - 0.3 

Entisols 2170 232 0.8 - 1.3 

Histosols 160 312 ? 

Inceptisols 950 324 8 - 13 

Mollisols 920 120 7 - 11 

Oxisols 1010 99 22 - 27 

Spososols 350 67 1 - 3 

Ultisols 1170 98 6 - 7 

Vertisols 320 18 1 - 2 

Gelisols 1120 238 0 

Others 1870 17 0.2 - 3 

Total 13050 1700 66 - 90 

Table 6.4 Estimates of historic SOC depletion from major soil orders. Source: Lal, 1999; Hillel and Rosenzweig, 2009. 

Erosion 

Water 

10 6 ha 

Area 

Wind 

10 6 ha 


Pg 

Light 343 269 2 - 3 

Moderate 527 254 10 - 16 

Strong 224 26 7 - 12 

Total 19 - 31 

Table 6.5 Estimates of historic SOC loss from accelerated erosion by water and wind. Source: Lal, 1999. 

and 19 to 31 Pg by erosional processes (Table 6.5). While the historic loss from Gelisols or permafrost soils is 

zero, these soils, which contain a vast amount of SOC reserves, are vulnerable to projected warming and the 

attendant positive feedback. 

Estimates of the historic C loss are useful as a reference point for assessing the technical potential of C resequestration 

in soil. While the loss of SOC can be rapid, especially in soils of the tropical ecosystems, the rate 

of re-carbonization is extremely slow. The slow rate of re-sequestration is a major challenge to identifying 

appropriate land use and to promoting adoption of soil/water/animal/plant management systems that could 

create a positive soil/ecosystem C budget. 


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6.2.7 | Future loss of SOC under climate change 

Projected changes in climate (temperature and precipitation) are likely to affect the SOC stock both directly 

and indirectly. Directly, the rate of decomposition by microbial processes is affected by both soil temperature 

and moisture regimes. Indirectly, changes in climate affect plant growth, net primary productivity, above 

and below-ground biomass, and the type and amount of residues with differential amounts of materials 

recalcitrance. Further, the rate and susceptibility to accelerated erosion, salinization and other degradation 

processes may be exacerbated by an increase in frequency of extreme events. Indeed, climate change can 

impact several soil forming factors, including rainfall, temperature, micro-organisms/biota and vegetation, 

thus affecting the rate of SOC accumulation (Jenny, 1930). Climate change may also alter species composition, 

and the rate of litter fall. However, disagreement exists regarding the effect of warming on SOC stock. 

The annual rate of litter return, on which the rate of SOC accretion depends, varies among biomes (White, 

1987; Grunwald, 1999). The rate of litterfall (Mg ha -1 yr -1 ) is estimated at 0.1 to 0.4 for alpine and arctic regions, 

2 - 4 for temperate grassland, 1.5 - 3 for coniferous forest, 1.5 - 4 for deciduous forest, 5 - 10 for tropical rainforest, 

and 1 to 2 for arable land (White, 1987). Increase in soil temperature may exponentially increase the rate of soil 

respiration (Tóth et al., 2007; Lenton and Huntingford, 2003). However, because of increase in the number and 

activity of soil fungi in the warmer soil, there may also be increase in the relative amount of lignin and other 

recalcitrant compounds (Simpson et al., 2007). The SOM decomposition is also more temperature-sensitive at 

low than at high temperature (Kirschbaum, 1995, 2000, 2006). 

Thus, knowledge about the temperature–sensitivity of diverse SOC fractions, and their change in the soil 

under climate change, is important. Change in temperature by 1º Celsius may decrease the turnover times of 4 -1 1 

percent and 8 -1 6 percent for the intermediate and stabilized fractions, respectively (Hakkenberg et al., 2008). 

The decomposition rate is also influenced by the presence of physicochemical protection mechanisms (Conant 

et al., 2011), especially occlusion within aggregates and by association with mineral surfaces (Freedman, 

2014). It is argued that CO 2 

emissions from soil response to climate warming are over-estimated, because the 

decomposition of old SOM is tolerant to temperature (Liski et al., 1999). Thus, the effects of warming on SOM 

decomposition are governed by complex and interactive factors, and are difficult to predict. Despite much 

research, no consensus has yet emerged on the temperature sensitivity of SOM decomposition (Davidson and 

Janssens, 2006). 


Global SOC stocks have been estimated at about 1 500 Pg C for the topmost 1 m. However, a large 

uncertainty attaches to this estimate, which cannot easily be assigned to a specific period in time. Local 

variations may also be high, for example for SOC stocks in arctic regions and peatlands. Estimates of SOC 

stocks below 1 m depth are still evolving, with a tendency for more recent estimates to be higher than previous 

values. Estimates of the historic loss of SOC pools are also highly variable, ranging from 40 to 537 Pg. The 

global loss of SOC pool since 1850 is estimated at about 66 ±12 Pg. The projected response of SOC stock to 

climate change is a debatable issue. While an increase in temperature may increase the rate of respiration at 

low soil temperature, it may also shift microbial populations to fungi, increase relative proportions of lignin 

and other recalcitrant fractions, and increase protective mechanisms such as aggregation and reaction with 

mineral surfaces. 


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6.3 | Soil contamination status and trends 


Soil contamination as a result of anthropogenic activities has been a widespread problem globally (Bundschuh 

et al., 2012; DEA, 2001; EEA, 2014; Luo et al., 2009; SSR, 2010). Soil contamination can be local or diffuse. Local 

soil contamination occurs where intensive industrial activities, inadequate waste disposal, mining, military 

activities or accidents introduce excessive amounts of contaminants. Diffuse soil contamination is the 

presence of a substance or agent in the soil as a result of human activity and emitted from dispersed sources. 

Diffuse contamination occurs where emission, transformation and dilution of contaminants in other media 

have occurred prior to their transfer to soil. The three major pathways responsible for the introduction of 

diffuse contaminants into soil are atmospheric deposition, agriculture, and flood events. These pathways can 

also cause local contamination in some instances. Causes of diffuse contamination tend to be dominated by 

excessive nutrient and pesticide applications, heavy metals, persistent organic pollutants and other inorganic 

contaminants. As a result, the relationship between the contaminant source and the level and spatial extent 

of soil contamination is indistinct. 

While some soil degradation processes are directly observable in the field (erosion, landslides, sealing or 

even decline of organic matter), soil contamination as well as soil compaction or decline in soil biodiversity 

cannot be directly assessed, which makes them an insidious hazard. Moreover, diffuse contamination is linked 

to many uncertainties. The diversity of contaminants (particularly of the persistent organic pollutants, which 

are in constant evolution due to agrochemical developments) and the transformation of organic compounds 

in soils by biological activity into diverse metabolites make soil surveys to identify contaminants difficult 

and expensive. The effects of soil contamination also depend on soil properties, as these have an impact 

on the mobility, bioavailability, residence time and levels of contaminants. Direct effects of pollutants may 

not be immediately revealed because of the capacity of soils to store, immobilize and degrade them. Effects 

can, however, suddenly emerge after changes such as changes in land use that may alter environmental 

conditions (Stigliani et al., 1991 - see also Chapter 7 on processes impacting service provision). Contaminants 

include inorganic compounds such as metallic trace-elements and radionuclides, and organic compounds 

like xenobiotic molecules. The application of some organic wastes to soils – for example, untreated biosolids 

– also increases the risk of spread of infectious diseases. A new challenge is that the so-called ‘chemicals of 

emerging concern’ (CECs) – for example, veterinary and human therapeutic agents such as antibiotics and 

hormones – are present in amendments added to soils, such as manures. These CECs can have an adverse 

effect on ecosystems and on human health (Jjemba, 2002; Osman, Rice and Codling, 2008; Jones and Graves, 

2010). 

6.3.2 | Global status of soil contamination 

In most developed countries, waste disposal and treatment, industrial and commercial activities, 

storage, transport spills on land, military operations, and nuclear operations are the key sources of local soil 

contamination. Management of local soil contamination requires surveys to seek out sites that are likely to be 

contaminated, site investigations where the actual extent of contamination and its environmental impacts 

are defined, and implementation of remedial and after-care measures. By contrast, diffuse soil contamination 

is much harder to manage: in many instances it is not directly apparent but it may cover very large areas and 

represent a substantial threat. Despite the fact that most developed countries have implemented long-term 

soil surveys, even these countries still lack a harmonized soil monitoring system, and the real extent of diffuse 

soil contamination is not known. 


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According to the most recent data provided by the European Environmental Agency (EEA, 2014), total 

potentially contaminated sites in Europe are estimated to be more than 2.5 million, of which 340 000 are 

thought to be actually contaminated. Approximately one third of the high risk sites have been positively 

identified as contaminated, and of these only 15 percent have so far been successfully remediated (EEA, 2014). 

While trends vary across Europe, it is clear that the remediation of contaminated sites is still a significant 

undertaking. Waste disposal and industrial activities are the most important sources of soil contamination 

overall in Europe. The most frequent contaminants are heavy metals and mineral oils (EEA, 2014). 

In the United States, sites contaminated with complex hazardous substances that impact soil, groundwater 

or surface water are placed on the Superfund National Priorities List (NPL). As of September 29, 2014, there 

were 1 322 final sites on the NPL. On 1 163 of these sites, measures to address the contamination threat 

have been completed. An additional 49 sites have been proposed. In addition, the Office of Solid Waste and 

Emergency Response (OSWER) has cleaned up over 540 000 sites and 9.3 million ha of contaminated land, 

all of which can be put back into use. In Canada, a total of 12 723 soil contaminated sites has been identified, 

with 1 699 sites related to surface soil contamination (Treasury Board of Canada Secretariat, 2014). The key soil 

contaminants include metals, petroleum hydrocarbons (PHCs), and polycyclic aromatic hydrocarbons (PAHs). 

The pattern of contamination in Australia is similar to that of other developed countries. Industry, including 

the petroleum industry, mineral mining, chemical manufacture and processing facilities, and agricultural 

activities with their use of P fertilizer and pesticides, have caused soil contamination with heavy metals, 

hydrocarbons, mineral salts, particulates, etc. The total number of contaminated sites is estimated at 80 000 

across Australia (DECA, 2010), with approximately 1 000 actual or potentially contaminated sites in South 

Australia (SKM, 2013). 

Developing countries are undergoing significant industrialization. If appropriate legal and regulatory 

frameworks and enforcement capability are not developed, this may lead to soil contamination and pose risks 

to the environment and human health. In large conurbations, there is also a need for adequate provision of 

sanitation and drainage so that household wastes are collected and managed safely. 

Asian countries experience considerable contamination of agricultural soil and crops by trace elements, and 

this contamination is becoming a threat to human health and the long-term sustainability of food production 

in the contaminated areas. In China, it is estimated that nearly 20 million ha of farmland (approximately one 

fifth of China’s total farmland) is contaminated by heavy metals (Wei and Chen, 2001). This may result in a 

reduction of more than 10 million tons of food supplies each year in China (Wei and Chen, 2001). Atmospheric 

deposition (mainly from mining, smelting and fly ash) and livestock manures are the main sources of trace 

elements contaminating arable soil (Luo et al., 2009). Among the different trace elements contaminating 

Chinese agricultural soils, Cd is the biggest concern. Due to its high mobility in the soil (except in poorly drained 

soil where sulphides are present), it can be easily transferred to the food chain and so poses risks to human 

health. Arsenic is also naturally present in groundwater in many regions of Southeast Asia. This represents 

a threat to agriculture, particularly in rice paddy fields where anaerobic conditions prevail (Smedley, 2003; 

Hugh and Ravenscroft, 2009). Asia is also the largest contributor to the atmosphere of anthropogenic Hg, 

which originates from the chemical industry, from Hg mining and from gold mining (Li et al., 2009). All across 

Asia, areas under rapid economic development are experiencing moderate to severe contamination by heavy 

metals (Ng, 2010). 

In many parts of Latin America, the results of anthropogenic activities, such as tailings and smelting 

operations in mining areas, have resulted in arsenic contamination in the soil. These operations enhance the 

mobilization of arsenic and cause adverse environmental impacts (see Section 4.3). Also in Latin America, the 

problem of arsenic contamination in water has been identified in 14 of the continent’s 20 countries: Argentina, 

Bolivia, Brazil, Chile, Colombia, Cuba, Ecuador, El Salvador, Guatemala, Honduras, Mexico, Nicaragua, Peru 


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and Uruguay. The number of exposed people in these countries is estimated to be about 14 million (Bundschuh 

et al., 2012; Castro de Esparza, 2006). It is also estimated that during the late 1980s and early 1990s, 3 000 to 

4 000 t of Hg were deposited in the Amazon basin as a result of artisanal gold-mining activities, mainly in 

Brazil, Bolivia, Venezuela and Ecuador (de Lacerda, 2003). In addition, intensive use of fertilizers and pesticides 

in many parts of Latin America contributes to soil contamination and causes a range of environmental 

pollution and human health problems (UNEP, 2010). 

In Africa, soil contamination has resulted from mining, spills, and improper handling of waste (Gzik et 

al., 2003; SSR, 2010; EA, 2010). The Nigerian federal government reported more than 7 000 spills between 

1970 and 2000. In Botswana and Mali, over 10 000 tonnes of pesticides, including DDT, aldrin, dieldrin, 

chlordane and heptachlor, have leaked from damaged containers and contaminated the soil (SSR, 2010). Soil 

contamination in the Near East and North Africa is linked to oil production and heavy mining. In arable land, a 

common source of soil pollution is the use of contaminated groundwater or wastewater for irrigation. 

6.3.3 | Trends and legislation 

In developed countries, legislation on contaminated land and the related regulatory mechanisms are well 

established. As a result, the extent of contaminated land is thoroughly reported. The European countries have 

created a common framework in the Thematic Strategy on Soil Protection (COM (2006) 231), which aims at 

sustainable use of soil, preservation of soil as a resource, and remediation of contaminated soil. The EC has 

also created networks such as CLARINET, NICOLE and SNOWMAN (Vicent, 2013). Investigations of suspected 

contaminated sites continue in Europe and as a result the total of contaminated sites listed is expected to 

increase by 50 percent by 2025 (EEA, 2007, 2012; EC, 2013). The number of remediated sites is expected to grow 

as well. In addition, regulation now requires industrial plants to control their wastes and prevent accidents, 

limiting the introduction of contaminants into the environment. As noted above, the United States has 

introduced a regulatory regime and has made significant progress on site clean-up. 

In Asia, early legislation on contaminated land management (CLM) focused on contamination of agricultural 

land caused by industrialization and urbanization. Thus Japan, Taiwan, Province of China and South Korea 

have developed comprehensive CLM frameworks of laws, regulations and guidelines. Other Asian countries, 

however, are still at early stages of developing a CLM framework (Ng, 2010). 

Atmospheric deposition (Section 4.4.1) is an important input of pollutants (Lofts et al., 2007) and air quality 

regulation to decrease the load of contaminants on soils is therefore important. In most developed countries, 

relevant legislation is well established. In the case of long-range atmospheric pollution, international 

agreements are needed. In this regard, the Convention on Long-Range Transboundary Air Pollution (LRTAP) 

was signed in 1979. Conceived in response to the detrimental impact of acid rain in Europe, the Convention 

entered into force in 1983. Over the past 30 years, the Convention has been extended by eight further protocols 

that target pollutants such as S, NOx, persistent organic pollutants, volatile organic compounds, ammonia 

and heavy metals. More recently, a global treaty to protect human health and the environment from the 

adverse effects of mercury - the 2013 Minamata Convention on Mercury - has been established. 

CECs require due attention and they can include, but are not limited to, nanoparticles, pharmaceuticals, 

personal care products, estrogen-like compounds, flame retardants, detergents, and some industrial 

chemicals (including those in products and packaging) with potential significant impact on human health 

and aquatic life (Jones and Graves, 2010). Electronic waste (also referred to as ‘e-waste’) is of great concern 

given the increasing volumes generated each year, the hazardous nature of some of the components, and the 

exportation of this waste from industrialized countries to recycling centres in China, India and Pakistan (UNEP 

DEWA/GRID-Europe, 2005). This chain risks violating the Basel Convention on the Control of Transboundary 

Movements of Hazardous Wastes and their Disposal, which was adopted in 1989 and came into force in 1992. 


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Recently some countries have implemented policies and programmes to encourage waste minimization. 

These programmes of ‘Extended Producer Responsibility’ make producers responsible for the costs of 

managing their products at the end of their life. This approach is expected to encourage the manufacture of 

more environmentally-friendly electronic products (UNEP DEWA/GRID-Europe, 2005). 

6.4 | Soil acidification status and trends 

6.4.1 | Processes and causes of acidification 

Soil acidity increases with the build-up of hydrogen (H+) and aluminium (Al 3 

+) cations in the soil or when 

base cations such as potassium (K+), calcium (Ca 2 

+), magnesium (Mg 2 

+) and sodium (Na+) are leached and 

replaced by hydrogen or aluminium (Bolan, Hedley and White, 1991; Helyar and Porter, 1989; von Uexküll and 

Mutert, 1995). The main causes of soil acidification are: (1) long term rainfall that results in on-site leaching 

of base cations; (2) draining of potentially acid sulphate soils; (3) acid deposition when urbanization, 

industrialization, mining, construction or dredging release acid substances into the air or water, causing offsite 

acidification; (4) excessive application of ammonium-based fertilizers (e.g. ammonium sulphate) as part 

of intensive agriculture cropping practices; and (5) deforestation and other land use practices that remove all 

harvested materials, often resulting in a drop of the pH in the topsoil. Only the first of these five causes is a 

natural phenomenon; all others are human-induced. 

In natural ecosystems, soils become more acid with time. Consequently old soils, particularly in humid 

climates or those developed from acidic rocks, are more weathered and acidic than younger soils or soils of dry 

climates or those developed from more basic rocks (Helyar and Porter, 1989; von Uexküll and Mutert, 1995). 

Soil acidification is of the greatest concern in soils that have a low capacity to buffer the decrease in pH and in 

soils that already have a low pH, such as acid soils in highly weathered tropical areas (Harter, 2007; Johnson, 

Turner and Kelly, 1982). Soil texture and soil organic matter content play an important role in the buffering 

capacity of a soil and hence in determining how prone a soil is to acidification (Helyar and Porter, 1989; Steiner 

et al., 2007). Light sandy soils poor in organic matter are the least buffered against acidification. 

Acid sulphate soils contain metal sulphides which, when exposed to oxidation, produce sulphuric acid. 

Inland, acid sulphate soils form naturally in aquatic ecosystems and also as a consequence of human-induced 

changes to land use and hydrology. Structures regulating water flow such as dams, weirs and locks prevent 

flushing of metals, salts and organic matter, and promote the build-up of acid sulphate soils. Acid sulphate 

soils also form in coastal areas and are common in mangrove forests, saltmarsh, floodplains, and salt- and 

freshwater wetlands (Lin and Melville, 1994; Pons, van Breemen and Driessen, 1982; Pannier, 1979). Due to the 

abundance of metal sulphides in rocks, mining activities also foster the formation of acid sulphate soils (Dent, 

1986). 

The atmospheric deposition of sulphur dioxide (SO 2 

), nitrogen oxides (NOx) and ammonia (NH 3 

) leads to 

acid deposition. This can affect not only areas near to the urban, industrial and mining sites where the oxides 

are produced and released into the environment, but also sites located far away (Fanning et al., 2004; Menz 

and Seip, 2004; Mylona, 1996; Orndorff and Daniels, 2004). The term ‘acid deposition’ includes both wet and 

dry (gaseous) precipitation, usually in the form of acid rain or fog. Besides affecting the chemistry of soil and 

water resources, acid deposition directly harms plants and fish. Acid deposition is currently a major concern in 

fast-developing countries such as China (Chen, 2007). 

Land use and soil management play a crucial role in determining the chemical characteristics of the soil. 

Intensive farming practices that employ large inputs of nitrogen fertilizers and remove large quantities of 


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products increase soil acidity (Barak et al., 1997; Bolan, Hedley and White, 1991). Indeed, the conversion of 

ammonium to nitrate releases hydrogen ions (H+) into the soil solution that can potentially lower the soil pH. 

This is a problem in soils with low ability to buffer the increase in H+ such as those poor in lime and negatively 

charged organic matter and clay. Harvesting has the potential to increase soil acidity by removing base cations 

from the soil. This is an issue in both agricultural and forested areas wherever large amounts of biomass are 

removed by crop harvesting and deforestation (Cavelier et al., 1999; von Uexküll and Mutert, 1995). 

6.4.2 | Impact of soil acidification 

On acid soils (pH < 5.5), crops and pastures suffer from the resulting increased phytotoxicity (Al, Fe, Mn, etc.), 

from the reduced availability of nutrients, and from decreased microbiological activity (Cronan and Grigal, 

1995; Robson and Abbott, 1989; Slattery and Hollier, 2002; Sverdrup and Warfvinge, 1993; Whitfield et al., 2010). 

Onsite soil acidification reduces net primary productivity and carbon sequestration by accelerating leaching 

of nutrients such as manganese, calcium, magnesium and potassium, resulting in nutrient deficiencies for 

plants (Haynes and Swift, 1986). On-site soil acidification is also responsible for the development of subsoil 

acidity (Tang, 2004), for the breakdown and subsequent loss of clay materials from the soil (Chen, 2007), 

and for the erosion which results from decreased groundcover (Slattery and Hollier, 2002). Soil acidification 

also leads to off-site effects such as surface water acidification through sediment losses, and groundwater 

enrichment of soluble metals. In turn, these processes mobilize heavy metals into water resources and the 

food chain (Driscoll et al., 2003; Reuss and Johnson, 1986; Schindler et al., 1980; Slattery and Hollier, 2002; 

Voegelin, Barmettler and Kretzschmar, 2003). 

6.4.3 | Responses to soil acidification 

Soil acidification is an insidious process. It develops slowly and, if not corrected by lime applications 

for example, can continue until the soil is irreparably damaged (Edmeades and Ridley, 2003; Liu and Hue, 

2001; Slattery and Hollier, 2002). Biological recovery can potentially be improved by an increase in pH and 

acid-neutralising capacity (ANC) (Marschner and Noble, 2000). Of main concern is subsoil acidity, which is 

particularly difficult to correct with conventional methods (Farina, Channon and Thibaud, 2000; Liu and Hue, 

2001; Hue and Licudine, 1999). Actions to mitigate global warming can reduce the emission of pollutants such 

as sulphur dioxide (SO 2 

) which contribute to soil acidification (NADP, 2014; Smith, Pitcher and Wigley, 2001; 

Vestreng et al., 2007). However, soil response to decreases in acid deposition is slow and acid-affected sites 

may require many decades to recover (Zhao et al., 2009). 

6.4.4 | Global status and trends of soil acidification 

Soil acidity is a serious constraint to food production worldwide. Traditionally it has been counteracted by 

applying lime to the topsoil but little could be done to increase the pH of the subsoil. Programmes to improve 

soil pH have been undertaken largely in developed countries, which are able to implement soil management 

plans to preserve soil properties and to bear the cost of lime to buffer soil acidity. However, even in developed 

counties, for example Australia, there have been cases where subsoil acidity increased due to failures in 

correcting topsoil acidity. In developing countries the situation is more stark as the use of lime is constrained 

by poverty. As a result, the farmed area affected by acidification is on the rise (Sumner and Noble, 2003). Soil 

acidification affects not only agricultural areas but also forests and grasslands. 

According to Sumner and Noble (2003), topsoil acidity (pH

Estimated dominant topsoil pH 

< 4.5 4.5 - 5.5 5.5 - 7.2 7.2 - 8.5 > 8.5 Water Rocks outcrops, glaciers, salt flat 

Figure 6.8 Estimated dominant topsoil pH. Source: FAO/IIASA/ISRIC/ISS-CAS/JRC, 2009. 

The main causes of soil acidification vary by region: 

• Regions where soil acidification occurs because of soil texture – parts of North America, Southeast, 

East and South Asia (Aherne and Posch, 2013; Eswaran et al., 1996; Hicks et al. 2008; Shamshuddin et 

al., 2014; Ouimet et al., 2006) 

• Regions where proximity to deltas and coastal plains is a primary cause – parts of West Africa (Bullock 

et al., 1996), 

• Regions where weather conditions are a main cause – parts of Africa and Asia (Breuning-Madsen and 

Awadzi, 2005; Drees, Manu and Wilding, 1993; Eswaran et al., 1996; Kottek et al., 2006; Wilke, Duke and 

Jimoh, 1984), 

• Regions where acid deposition is an important factor – parts of East Asia and North America (Aherne 

and Posch, 2013; Quinn, 1989; Wolt, 1981) 

• Regions where the massive application of ammonium-based fertilizers plays an important role – parts 

of East and South Asia (Guo et al., 2010; Wang, Zhang and Zhang, 2010). 

In Europe, soil acidification is an issue only in some highly urbanised and industrialized hotspots (EEA, 

2010; Kopáček et al., 2004; Menz and Seip, 2004; Moser and Hohensinn, 1983). In the Southwest Pacific, soil 

acidification is of concern only in intensively farmed areas (Brennan, Bolland and Bowden, 2004; Hartemink, 

1998; Xu et al., 2002; Lockwood et al., 2003; NLWRA, 2001). Thus soil acidification affects all regions to some 

extent, but it is of main concern in poor and developing countries which are growing rapidly but are unable to 

buffer the decrease in soil pH through conventional means. 

6.5 | Global status of soil salinization and sodification 

6.5.1 | Status and extent 

Salt-affected soils occur in more than 100 countries and their worldwide extent is estimated at about 1 

billion ha. Salt-affected soils include those affected by salinity, where the electrical conductivity of the soil 

exceeds 4dSm–1; and those affected by sodicity, where the exchangeable sodium percentage (ESP) exceeds 6 

(Ghassemi, Jakeman and Nix, 1995). Saline soils contain excessive soluble salts, mainly sodium chloride (NaCl) 

and sodium sulphate (Na 2 

SO 4 

) or other neutral salts. These salts increase osmotic pressure, diminish water 


124

availability and inhibit plant growth. Sodic soils generally have a low salt content but a high ESP, which causes 

dispersion of clay particles and results in deterioration of the soil structure. These soils generally have low air 

and water permeability and a pH above 8.2. 

Salinity problems are encountered in all climates and are a consequence of both natural (primary) and 

human-induced (secondary) processes. Soil salinity and sodicity problems are more common where rainfall 

is insufficient to leach salts and excess sodium ions out of the rhizosphere. Salt-affected soils often occur 

on irrigated lands, especially in arid and semiarid regions, where annual rainfall is insufficient to meet the 

evapotranspiration needs of plants and to provide for leaching of salt. In humid areas, soluble salts are carried 

down through the soil profile by percolating rainwater and ultimately are transported to sea. 

Although salt-affected soils are widespread and an increasingly severe problem, no accurate recent 

statistics are available on their global extent. The best available estimates suggest that about 412 million ha are 

affected by salinity and 618 million ha by sodicity (UNEP, 1992), but this figure does not distinguish areas where 

salinity and sodicity occur together. The Soil Map of the World (FAO/UNESCO, 1980) depicted a similar extent 

Continent 

Saline soils 

(million ha) 

Sodic soils 

(million ha) 

Total 

(million ha) 

Africa 122.9 86.7 209.6 

South Asia 82.3 1.8 84.1 

North and Central Asia 91.5 120.2 211.7 

Southeast Asia 20.0 - 20.0 

South America 69.5 59.8 129.3 

North America 6.2 9.6 15.8 

Mexico/Central America 2.0 - 2.0 

Australasia 17.6 340.0 357.6 

World total 412.0 618.0 1030 

Table 6.6 Distribution of salt-affected soils in drylands different continents of the world. Source: UNEP, 1992. 

of 953 Mha affected by salinity (352 million ha) and sodicity (580 million ha). Table 6.6 shows the distribution 

of dryland salinity in different continents. 

Human-induced salinity, mainly caused by irrigation without adequate drainage, affects a much smaller 

area than natural salinity. According to GLASOD, the extent of human-induced salinity is about 76 million ha 

(Oldeman, Hakkeling and Sombroek, 1991) of which 52.7 million ha occurs in Asia. In Europe, significant parts 

of Spain and areas in Italy, Hungary, Greece, Portugal, France and Slovakia are also affected by human-induced 

salinization. 

In 2006 the global area equipped for irrigation stood at 301 million ha. At present in developing countries, 

irrigated agriculture covers about one fifth of all arable land, but accounts for nearly half of all crop production 

and 60 percent of cereal production. About 70 percent of the world area equipped for irrigation is in Asia where 

it accounts for 39 percent of the cultivated area. India and China each have 62 million ha equipped for irrigation 

(FAO, 2011). An estimated 60 million ha (or 20 percent of the total irrigated area) are affected by soil salinity, 

of which 35 million ha are located in four countries e.g. Pakistan (3.2 million ha), India (20 million ha), China (7 

million ha) and the United States (5.2 million ha). Other countries with large amounts of salt-affected lands 

in irrigation districts include Afghanistan, Egypt, Iraq, Kazakhstan, Turkmenistan, Mexico, Syria and Turkey 

(Squires and Glenn, 2011). 

Australia has the largest extent of naturally sodic soils of any continent (Table 6.5). 


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6.5.2 | Causes of soil salinity 

The distribution of salt-affected soils varies geographically with climate, landscape type, agricultural 

activities, irrigation methods and policies related to land management. 

Natural causes of salinity and sources of salt 

1. Rock weathering: Significant quantities of sodium, and to a lesser extent chloride, occur widely in the 

parent rocks from which soils form. Over time, rock weathering can lead to appreciable salt accumulation 

in soils if leaching is restricted. Rock weathering is the primary source of salt in seawater. 

2. Sea water and accession of salt in marine sediments: Saline soils can form from sediments and parent materials 

that were once under the sea. Likewise, the salts can be due to tidal inundation. Typical examples include 

the pseudo-delta of Senegal and the Gambia and in the Philippines where coastal tideland reclamation 

has created about 0.4 million ha of agricultural salt-affected soils. In the United Arab Emirates, areas 

along the coastal sabkha (salt marshes or lagoonal deposits) are highly salinized (28.8 dS m -1 ). In the 

coastal region of the Abu Dhabi Emirate, salinity is more than 200 dS m -1 (Abdelfattah and Shahid, 2007) 

3. Atmospheric deposition: Salt derived from the sea, either deposited via rain or dry fallout, is the primary 

source of salt across large areas: for example, many millions of hectares in southern Australia. In arid 

areas, salt can also be derived from dry lake beds and then blown considerable distances by wind (e.g. 

Eurasia and parts of Australia). 

Human-induced causes 

1. The management of land and water resources is responsible for the development of human-induced 

saline and sodic soils. The main causes are: 

2. Poor drainage facilities which induce a rise of the groundwater table. This is a major cause of soil 

salinization in India, Pakistan, China, Kenya and the Central Asian countries. 

3. The use of brackish groundwater for irrigation. This is a major cause of secondary salinization in parts 

of Asia, Europe and Africa. 

4. The intrusion of seawater in coastal areas, for example in Bangladesh. 

5. Poor on-farm water management and cultural practices in irrigated agriculture. 

6. Continuous irrigation over very long periods, particularly in the Middle East. 

7. Replacement of deep rooted perennial vegetation with shallower rooted annual crops and pastures 

that use less water leading to the rise of saline groundwater, for example southern Australia. 

6.5.4 | Trends and impacts 

Soil salinity is becoming a significant problem worldwide. From the very scattered information on the 

extent and characteristics of salt-affected soils, salinity and sodicity are rapidly increasing in many regions, 

both in irrigated and non-irrigated areas. Increasing soil salinity problems are taking an estimated 0.3 to 1.5 

million ha of farmland out of production each year and decreasing the production potential of another 20 to 

46 million ha. The annual cost of salt-induced land degradation was estimated in 1990 at US$ 264 ha−1. By 

2013, the inflation-adjusted cost of salt-induced land degradation was reported as US$ 441 ha–1 (Qadir et al., 

2014). 

6.5.5 | Responses 

There are many available responses to contain the salinity threat. These include: (1) direct leaching of salts; 

(2) planting salt tolerant varieties; (3) domestication of native wild halophytes for use in agro-pastoral systems; 

(4) phytoremediation (bioremediation); (5) chemical amelioration; and (6) the use of organic amendments. 


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In several Asian countries, a blend of engineering, reclamation and biological approaches has been adopted 

to address salinity and waterlogging problems. In Pakistan, engineering solutions included large-scale Salinity 

Control and Reclamation Projects (SCARPs), which covered 8 million ha at an estimated cost of US$2 billion 

(Qureshi et al., 2008). Two big drainage water disposal projects were also undertaken. Measures to address the 

saline soil problem included leaching of salts by excess irrigation, use of chemicals (such as gypsum and acids), 

the addition of organic matter, and biological measures such as salt-tolerant plants, grasses, and shrubs. 

Improvements in on-farm water and crop management have also been practiced. In North America, 

changes in land use and management practices have reduced the risk of salinization and helped to improve 

soil health and agri-environmental sustainability. 

In Iraq and Egypt, surface and subsurface drainage systems have been installed to control rising water 

tables and arrest soil salinity. In Iran, Syria and other Gulf countries, crop-based management, and fertilizers 

are used to combat salinization (Qadir, Qureshi and Cheraghi, 2007). In Iran, Haloxylon aphyllum, Haloxylon 

persicum, Petropyrum euphratica and Tamarix aphylla are potential species for saline environments 

(Djavanshir, Dasmalchi and Emararty, 1996). Also in Iran, Atriplex has been shown to be a potential fodder 

shrub in the arid lands which could bring annual income as high as US$ 200 ha -1 (Koocheki, 2000; Nejad and 

Koocheki, 2000). Breeding of salt tolerant crop varieties (e.g. wheat, barley, alfalfa, sorghum etc.) is also a 

recognized management response for saline environments. However, most results have been obtained in 

controlled environments, with few real field results so far. 

The use of organic amendments in Egypt showed that the mixed application of farmyard manure and gypsum 

(1:1) significantly reduces soil salinity and sodicity (Abd Elrahman et al., 2012). Recently, phytoremediation or 

plant based reclamation has been introduced in the Near East region. In Sudan good responses for control 

of sodicity have been obtained through phytoremediation. The production of H+ proton in the rhizosphere 

during N-fixation from legumes such as the hyacinth bean (Dolichos lablab L.) removed as much Na+ as 

gypsum application. This indicates the importance of this technology in calcite dissolution of calcareous salt 

affected soils (Mubarak and Nortcliff, 2010). 

6.6 | Soil biodiversity status and trends 


Over the last few decades the importance of soil biota for terrestrial functioning and ecosystem services has 

emerged as an important focus for soil science research. Current evidence shows that soil biota constitute an 

important living community in the soil system, providing a wide range of essential services for the sustainable 

functioning of global terrestrial ecosystems and thereby impacting human wellbeing, directly and indirectly 

(van der Putten et al., 2004). Soil organisms (e.g. bacteria, fungi, protozoa, insects, worms, other invertebrates 

and mammals) shape the metabolic capacity of terrestrial ecosystems and many soil functions. Below-ground 

biodiversity represents one of the largest reservoirs of biodiversity on earth (Bardgett and van der Putten, 

2014). Essential services provided by soil biota include: regulating nutrient cycles; controlling the dynamics of 

soil organic matter; supporting soil carbon sequestration; regulating greenhouse gas emissions; modifying 

soil physical structure and soil water regimes; enhancing the amount and efficiency of nutrient acquisition 

by vegetation through symbiotic associations and nitrogen fixation by bacteria; and influencing plant and 

animal health through the interaction of pathogens and pests with their natural predators and parasites. 

Fungi and bacteria are important decomposers in the soil. They are remarkably efficient. The smaller the 

pieces to be decomposed, the faster these microorganisms are able to do their job. Organic waste such as 


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leaf matter and the droppings of herbivores first feed a host of small animals including insects, earthworms 

and other small invertebrates which live in the plant litter. The combined fauna break up the organic matter, 

digesting part of it, and thus facilitating the task of the microorganisms and invertebrates that complete 

the process of decomposition. In turn, soil macro-fauna affect soil organic matter dynamics through organic 

matter incorporation, decomposition and the formation of stable aggregates that protect organic matter 

against rapid decomposition. Successive decomposition of dead material and modified organic matter results 

in the formation of a more complex organic matter called humus, which affects soil properties by increasing 

soil aggregation and aggregate stability, increasing the cation-exchange capacity (the ability to attract and 

retain nutrients), and increasing the availability of N, P and other nutrients. 

Many scientists have reported the role of macro-fauna in the accumulation of soil organic matter. For 

example the work by Snyder, Baas and Hendrix (2009), showed that millipedes and earthworms, both by 

themselves and taken together, reduce particulate organic matter. In addition, earthworms create significant 

shifts in soil aggregates from the 2000–250 and 250–53 µm fractions to the > 2000 µm size class. Earthworminduced 

soil aggregation was lessened in the 0-2 cm layer in the presence of millipedes. Further, Hoeksema, 

Lussenhop and Teeri (2000) found that in high-N soil with twice-ambient CO 2 

there was a higher density of 

predator/omnivores, lower diversity, and a larger value of Bonger’s Maturity Index compared to ambient CO 2 

. 

In this experiment, fine root biomass and turnover were significantly greater under elevated CO 2 

. This indicates 

higher vigour in plant root development and growth and hence increased carbon sequestration conditioned 

by enhanced soil biota activity. 

Studies also show the role of soil biota (including fungi, bacteria and plant parasitic nematodes) as 

pathogens and parasites or herbivores in decreasing root and plant productivity or reducing fruit quality. 

Recent research has focussed on the use of nematode and fungal resistant plant species or of other soil 

organisms as suppressive agents to modify the pathogens. 

6.6.2 | Soil biota and land use 

Losses in soil biodiversity have been demonstrated to affect multiple ecosystem functions including plant 

diversity, decomposition, nutrient retention and nutrient cycling (Wagg et al., 2014). Links between aboveground 

and below-ground communities (Wardle et al., 2004; De Deyn and van der Putten, 2005; Bardgett 

and van der Putten, 2014) suggest that factors affecting above-ground extinction may also be affecting soil 

organisms. 

Agricultural intensification, in particular, may reduce soil biodiversity, leading to decreased food-web 

complexity and fewer functional groups (Tsiafouli et al., 2015). Other driving forces that influence biodiversity 

in agricultural soils include the influence of crops/plants, fertilizers and pH, tillage practices, crop residue 

retention, pesticides, herbicides and pollution (Breure et al., 2004; Bardgett and van der Putten, 2014). Soil 

biological and physical properties (e.g. temperature, pH, and water-holding characteristics) and microhabitat 

are altered when natural habitat is converted to agricultural production (Crossley et al., 1992; Bardgett and 

van der Putten, 2014). Changes in these soil properties may be reflected in the distribution and diversity of soil 

meso fauna. Organisms adapted to high levels of physical disturbance become dominant within agricultural 

communities, thereby reducing richness and diversity of soil fauna (Paoletti et al., 1993). 

The management practices used in many agro-ecosystems (e.g. monocultures, extensive use of tillage, 

chemical inputs) degrade the fragile web of community interactions between pests and their natural enemies. 

The intensification of agricultural management may result in increased incidence of pests and diseases, 

with numerous studies reporting declines in the biodiversity of soil fauna (Decaens and Jimenez, 2002; 

Eggleton et al., 2002). In addition, the contribution of soil fauna globally to organic matter decomposition 

rates may be highly dependent on the temperature and moisture of an ecosystem (Wall et al., 2008). This 


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underlines the need for global-scale assessments. In a global study of soil fungi using 365 soil samples from 

natural ecosystems, Tedersoo et al. (2014) found that distance from the equator and annual precipitation had 

considerable effect on fungal species richness. They also identified various other controls on soil fungi and this 

is starting to provide a benchmark for assessing the impacts of human activities on an important component 

of soil biodiversity. 

Soil management strongly influences soil biodiversity, resulting in changes in abundance of individual 

species. Using a soil biodiversity pressure index calculation from the European Soil Data Centre, Gardi, 

Jeffery and Saltelli (2013) estimated that 56 percent of soils within the European Union have some degree of 

threat to soil biodiversity. Based on a questionnaire completed by 20 experts, the study found that the main 

anthropogenic pressures on soil biodiversity are (in order of importance): (1) intensive human exploitation; (2) 

reduced soil organic matter; (3) habitat disturbance; (4) soil sealing; (5) soil pollution; (6) land-use change; (7) 

soil compaction; (8) soil erosion; (9) habitat fragmentation; (10) climate change; (11) invasive species; and (12) 

GMO pollution (Gardi, Jeffery and Saltelli, 2013). 

There is some experimental evidence that there may be threshold levels of soil biodiversity below which 

functions decline (e.g. Van der Heijden et al., 1998; Liiri et al., 2002; Setälä and McLean, 2004). However, in 

many instances this is at experimentally prescribed levels of diversity that rarely prevail in nature. Although 

some studies demonstrate some functional redundancy in soil communities (e.g. Setälä, Berg and Jones, 

2005), high biodiversity within trophic groups may be advantageous since the group is likely to function more 

efficiently under a variety of environmental circumstances, due to an inherently wider potential. In a synthesis 

of diversity-function relationships of soil biodiversity focusing on carbon cycling, Nielsen et al., (2011) concluded 

that although there is considerable functional redundancy in soil communities for general processes, change 

may readily have an impact on specialized processes. However, data to support this conclusion are still limited. 

More diverse systems may be more resilient to perturbation since, if a proportion of components are removed 

or compromised in some way, others that prevail will be able to compensate (Kibblewhite, Ritz and Swift, 

2008). 


A comprehensive global-assessment on below-ground biodiversity has yet to be carried out. Although there 

is a Global Soil Biodiversity Atlas (EU/JRC, in press), no benchmark values exist on a global scale. This makes it 

difficult to quantify changes or future losses that may result from natural or anthropogenic-induced changes. 

Although progress is being made, few monitoring programs exist that quantify soil biodiversity across regions 

and at multiple trophic levels, especially outside of Europe. Regarding the threats to soil biodiversity and the 

effects on ecosystem functioning, more comparative and coordinated studies (from local to global scales) are 

needed across all ecosystems. These studies should quantify threats and determine the consequences of soil 

biodiversity loss to ecosystem functions, as well as the effects of interactions between threats. In addition, 

there is a need for standardization of methods in soil biodiversity studies so that multiple datasets can be 

synthesized and benchmark values for global soil biodiversity may be established. The use of DNA-based 

approaches is accelerating the speed at which data is being collected for all organisms. However, although 

sequencing data must be deposited into a public database (e.g. Genbank) before publication, the majority of 

morphological data still remains inaccessible and hence largely unavailable for meta-analysis. International 

initiatives such as the Global Soil Biodiversity Initiative 2 , ECOFINDERS, and the EU-sponsored Global Soil 

Biodiversity Atlas are steps in the right direction but a common database of soil biodiversity data for both 

morphological and molecular data is still needed (Orgiazzi et al., 2015). 

2 www.globalsoilbiodiversity.org 


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Figure 6.9 Historical and predicted shift of the urban/rural population ratio. Source: UN, 2008. 

6.7 | Soil sealing: status and trends 

For millennia, the vast majority of people lived a rural life, largely dependent on agriculture and other rural 

occupations. Only over the last two centuries has the ratio between the urban and non-urban population 

started to change rapidly. In 1800, only 3 percent of the world’s population lived in cities; in 1900 14 percent, 

47 percent in 2000, 50 percent in 2007, and 54 percent in 2014. The proportion of the urban population is 

expected to rise to 66 percent by 2050 (Figure 6.9). 

The world’s urban population is growing and cities are expanding in order to accommodate the increasing 

population and economic activity. It is not known with any certainty what share of the Earth’s land surface (ca. 

144 million km 2 ) is now occupied by cities or how much land will be required to accommodate the expected urban 

expansion (Potere et al., 2009). One of the most accurate estimates of the extent of urban areas at global scale, 

based on the use of MODIS satellite data at a resolution of 500 m, indicates for the year 2000 an area of 657 000 km 2 

(Potere et al., 2009), equivalent to 0.45 percent of the Earth’s land surface. Urbanization is an important contributor 

to regional and global environmental change (Foley et al., 2005, Ellis and Ramankutty, 2008). The growth of cities has 

a vast impact on the landscape and significant impact on soil resources (Chen, 2007; Gardi et al., 2014). 


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Figure 6.10 Urbanisation of the best agricultural soils. 

Between 1990 and 2000, the total extent of urban areas worldwide increased by 58 000 km 2 . During this period, 

2.8 percent of Europe’s total land was affected by land use change, including a significant increase in urban land. Of 

the total land take in the EU between 1990 and 2000, 71 percent was for agriculture. Between 2000 and 2006, the 

equivalent figure was only 53 percent. Had the land taken for urban expansion been devoted to agriculture instead, 

the land would have produced more than 6 Mt of wheat. More generally, the best quality soil in alluvial plains is often 

sealed by expanding cities and the rate of conversion is expected to increase rapidly, especially in developing countries. 


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The term ‘soil sealing’ is defined as the permanent covering of the soil surface with an impermeable material. 

Urbanisation affects the inner urban ecosystem as well as the neighbouring ecosystems. Besides the economic 

and social effects, negative environmental effects are predominantly linked to land consumption, the loss of high 

quality agricultural soil (Figure 6.10), the destruction of habitat, fragmentation of existing ecosystems, increased fuel 

consumption, air, water and soil pollution, and the alteration of microclimate. 

Soil sealing is in practice equivalent to total soil loss – virtually all services and functions are lost except 

the carrying capacity as a platform for supporting infrastructure. The main negative impacts on ecosystem 

services include: virtually total loss of food and fibre production; a significant decrease or total loss of the soil’s 

water retention, neutralization and purification capacities; the loss of the carbon sequestration capacity; and 

a significant decrease in the ability to provide (micro) climate regulation. The results include the loss of habitat 

for soil organisms, loss of soil biodiversity and nutrient cycling, and often a diminished landscape and natural 

heritage. 

Urban expansion is, of course, both beneficial and essential. Historically, the beginning of the most 

important civilizations was associated with both the development of agriculture and the creation of urban 

settlements. As early as 3000 BC, cities had arisen in the Fertile Crescent, on the banks of Nile, in the Indus 

River valley and along major rivers in China. However, the very rapid urban expansion of recent times is creating 

the need for trade-offs, including decisions regarding soil health and the rate of soil sealing. 

6.8 | Soil nutrient balance changes: status and trends 


Though changes in soil nutrient balances may possibly affect all types of terrestrial ecosystems, rapid 

changes are more likely to occur in managed ecosystems as a result of the export of biomass or the addition 

of nutrients to sustain productivity. These managed ecosystems include cropland, intensively or extensively 

grazed rangelands or meadows, and forests. Monitoring changes in soil nutrient content is of particular 

relevance in managed ecosystems because it provides a means to evaluate future changes in the ability of soils 

to maintain their ecosystemic functions. On the one hand, negative balances (‘nutrient mining’) ultimately 

translate into crop nutrient deficiencies (, food production deficits and human nutritional imbalances. On the 

other hand, positive balances may lead to negative environmental and health externalities. Eutrophication, 

increased frequency and severity of algal blooms, hypoxia and fish kills and loss of habitat and biodiversity 

have been related to excessive inputs of N and P into fresh and coastal waters. Excess application of N has 

also lead to widespread contamination of groundwater by NO 3 

. Gaseous emissions of ammonia and nitrous 

oxide may also degrade air quality and contribute to acidification, eutrophication, ground-level ozone and 

climate change (Oenema, 2004; Chadwick et al., 2011). In addition, strongly positive balances may reflect 

poor economic management of managed ecosystems. Nutrient balances can thus be viewed as indicators of 

sustainability of human-induced land use changes and land use practices. 

Soil nutrients include the macronutrients nitrogen (N), phosphorus (P), potassium (K), calcium (Ca), 

magnesium (Mg) and sulphur (S). In addition, the soil supplies micronutrients (boron, copper, iron, 

manganese, chloride, molybdenum, zinc), whose concentrations in plants are typically one or two orders of 

magnitude less than those of macronutrients. In most cases, N, P and K taken individually or in combination 

are the most limiting nutrients for plant growth. This section will therefore focus on these three elements. In 

soils, these nutrients may be present in different pools. Because the amount of nutrients in certain pools may 

vary strongly and erratically over short time intervals, stocks and mass balances are generally calculated on 

the basis of total nutrient content, without distinction among different forms (Roy et al., 2003). 


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Figure 6.11 Major components of the soil nutrient balance. 

The red discontinuous line marks the soil volume over which the mass balance is calculated. Green arrows correspond to inputs and 

red arrows to losses. ΔS represents the change in nutrient stock. 

6.8.2 | Principles and components of soil nutrient balance calculations 

Because the magnitude of the nutrient fluxes is often small compared to the total stock of nutrients in the 

soil profile, changes in soil nutrient stocks can be rather slow and difficult to detect over short time scales 

(< decades). Hence calculating nutrient balances from nutrient flows rather than from changes in nutrient 

stocks has been preferred in many studies (Figure 6.11). 

Table 6.6 lists the main inputs and outputs used for calculating the mass balances of N, P and K. Inorganic 

amendments are mostly composed of mineral fertilizers, but also comprise urine or minerals contained in 

irrigation water. Organic amendments include liquid, semi-solid or solid manures, compost, mulching 

material not produced on-site, and household refuse. It also includes faeces dropped by animals. In systems 

such as urban gardening, the re-use of waste water may also input organic compounds. Biological fixation by 

bacteria is restricted to N. Wet deposition refers to nutrients supplied with rainwater, whereas dry deposition 

refers to nutrients deposited as dust and aerosols. Dry deposition is a particularly important phenomenon 

in the case of K in areas downwind of major dust producing areas (e.g. West Africa;). Sedimentation refers to 

the deposition of sediment eroded upstream or to sediment deposited during river flooding. Additional fluxes 

may exist in specific situations (e.g. nutrients in subsurface lateral flows ; supply of NO 3 

from groundwater (). 


133

N P K 

Nutrient inputs 

Inorganic amendments Yes Yes Yes 

Organic amendments Yes Yes Yes 

Biological fixation Yes No No 

Dry or wet deposition Yes Yes Yes 

Sedimentation and run-on Yes Yes Yes 

Nutrient outputs 

Harvested products Yes Yes Yes 

Grazed products Yes Yes Yes 

Leaching Yes Generally negligible Low 

Gaseous emissions Yes No No 

Erosion and runoff Yes Yes Yes 

Table 6.7 Major components of soil nutrient mass balances for N, P and K 

The main losses are related to nutrients contained in exported harvested products (crops or fodder), and 

nutrients contained in food ingested by primary grazers (Table 6.7). Nutrients may also be lost by gaseous 

emissions (NH 3 

, N 2 

, N 2 

O), through erosion and in surface runoff, or by leaching. The latter applies mostly to 

NO 3 

-N, to a lesser extent to NH 4 

-N and K, and to a very limited extent to PO 4 

-P except in coarse textured soils 

saturated with P. 

6.8.3 | Nutrient budgets: a matter of spatial scale 

The larger the spatial scale, the more certain nutrient flows are internalized. For instance, in a selfsufficient, 

well-managed farm, the net balance may be nil or close to nil. However, different parts of the farm 

may well have very different balances. Likewise, in extensively-managed agropastoral systems, nutrient flows 

mediated through livestock occur between rangelands and croplands. At a regional scale, the balances may 

thus be nil or only slightly negative, whereas large imbalances exist within the region (see Box 6.1). 

At the global scale, fertilizer use and the growing of leguminous crops have resulted in a doubling of the 

rate at which N enters the terrestrial ecosystems as compared to pre-industrial levels. Likewise, the use of P 

fertilizers, animal feed supplements and detergents has led to a doubling of P inputs in the environment as 

compared to background P release from weathering. This is indicative of a net positive balance but hides large 

regional disparities. Bouwman, Beusen and Billen (2009) calculated global soil N and P balances for the year 

2000. Outputs were restricted to harvested and grazed crops and grasses, whereas inputs included manure, 

fertilizers, N deposition and N fixation. These authors estimated the inputs to soils at 249 Tg N and 31 Tg P yr -1 and 

losses through harvest and grazing at 93 Tg N and 16 Tg P yr -1 . Assuming no build-up of N in the soil, their model 

predicted that 16 percent (41 Tg yr -1 ) of the inputs may be lost by erosion and leaching, thereby contributing 

to a loss in environmental quality. In the case of P, their calculations predicted a net accumulation of P at a 

rate of 12 Tg yr -1 and losses of P through leaching and erosion of 2 Tg yr -1 . On a continental scale, considering 

both natural and agro-ecosystems, balances were always positive and comprised between 8.5 (North Asia) 

and 35 (Europe) kg N ha -1 yr -1 , and between 0.22 (Africa) and 5.5 (Europe) kg P ha -1 yr -1 . Focusing specifically on 

P and cropland, but restricting the balance calculations to fertilizer and manure inputs and harvest outputs, 

highlighted large P deficits in South America, northern United States and eastern Europe. Large P surpluses 

were found in the coastal United States, western and southern Europe, East Asia and southern Brazil. 


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Within the same continent, large variations in nutrient balances may occur. For 13 African countries, 

estimated balanced or negative nutrient budgets for N, P and K. At the national level, estimated soil nutrient 

balances for the year 2000 ranged from -2 to -60 kg N ha -1 yr -1 , from 0 to -1 1 kg P ha -1 yr -1 , and from -2 to -61 

kg K ha -1 yr -1 . A later study at 1 km² resolution confirmed the overall negative balances but highlighted larger 

variability over short distances. The rate of nutrient mining by crops was generally low or moderate, because 

of low land productivity (low yields), but accumulated over many decennia nutrient depletion may become 

severe and may be strongly aggravated by soil erosion. 

Based on a review of 57 nutrient budget studies related to the African continent, confirmed that N budgets 

at field and farm scale were largely negative whereas for phosphorus negative balances were reported in only 

56 percent of the studies. Going from the continental scale to the plot scale, there was a tendency for the 

variability in nutrient budgets to increase. This is to be expected, as land uses and management practices in 

smallholder agriculture in Africa are highly diversified between farms, within farms and even within plots. The 

study did not find a clear trend in the magnitude of the nutrient budgets from plot to continental scales. This 

is in contrast to other studies which did report increasingly negative balances as the scale increased. 

Box 6.1 | Livestock-related budgets within village territories in Western Niger 

(Schlecht et al., 2004) 

In the Sahelian zone of West Africa, between 1.5 and 9 kg N ha -1 yr -1 and between 0.06 and 0.7 kg P ha -1 

yr -1 are taken in by grazing livestock. The quantity varies by location and land use type (rangeland, 

cropland, fallow). However, up to 95 percent of the nutrients consumed by livestock are recycled through 

faeces. About 40-50 percent of these faeces end up being spatially concentrated at corralling spots or 

in farmyards, which represent only a few percent of the total village lands. Though nutrient in- and outflows 

related to livestock account for only a small fraction of the nutrient flows in Sahelian crop-livestock 

systems, livestock thus plays a major role in the spatial redistribution of nutrients. Negative balances 

occur on rangelands and variable (positive or negative) balances are found in croplands depending on the 

intensity of application of organic amendments. 

At even smaller scales, differences in soil fertility may arise from differential nutrient budgets. Strong 

gradients in soil fertility have been reported around villages, compounds, trees and shrubs as a result of 

higher levels of inputs (litter, household refuse, human excreta, manure and urine from resting animals, 

sedimentation, etc.) near these features. These are referred to as ‘fertility rings’ or ‘fertility islands’. 

6.8.4 | Nutrient budgets: a matter of land use system, land use type, managementand 

household equity 

Nutrient balances vary greatly across land use (LU) systems. Intensive growing of industrial crops in Europe 

is generally characterized by excess inputs of N, despite a recent tendency towards reduced fertilization 

driven by EU regulations and the economics of fertilizer use. As a result of the decoupling of livestock and land 

and because livestock are increasingly fed with imported feed, pastures are commonly exposed to excessive 

applications of manure (e.g. in Normandy in France, and in Denmark and Holland). Regarding P, after decades 

of excess application of P, there is nowadays a tendency for farmers to reduce their P application rates, or even 

to stop applying P altogether and to rely only on accumulated soil reserves and P released from soil mineral 

weathering. 

At the other extreme, subsistence farming in developing countries is commonly characterized by negative 

balances, reflecting nutrient mining (Roy et al., 2003). examined nutrient balances for different land uses in 


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a Kenyan district. N deficits in excess of -1 00 kg ha -1 yr -1 were found for maize, sugar cane, and pyrethrum. P 

deficits in excess of -1 0 kg ha -1 yr -1 were found for sugar cane, pyrethrum, and beans, but P excesses occurred 

in tea and maize-bean plots. Except for coffee, tea and seasonal fallow, K deficits in excess of -50 kg ha -1 yr -1 

occurred in all systems. These observed differences reflect differences in the use of (in-)organic amendments, 

but also nutrient transfers across LU types. In the case of coffee for instance, mulching is recommended, which 

is done by using residues from other crops (e.g. bananas) or grasses from fallow land. 

In Asia, both strongly positive and strongly negative balances have been reported. K deficits have been 

reported for rice-based systems across several Asian countries ranging from -25 to -70 kg ha -1 yr -1 . also reported 

K deficits in 71 paddy farms in south China, but found N and P surpluses. Based on negative nutrient balances 

for Bangladesh, Vietnam, Indonesia, Myanmar, the Philippines, and Thailand, and positive balances for Japan, 

Malaysia and Korea, it has been argued that lower-income countries with large and growing population were 

more likely to present negative balances whereas higher income countries with stable populations tended to 

have positive balances. In sub-Saharan Africa, the larger the population density, the more negative the N and 

P balances. 

For similar systems, differences in nutrient balances may also arise from variable access of farmers to 

external inputs. In the Sudanian zone of west Africa, cultivated plots near hamlets tended to have less negative 

or more positive balances than plots near larger villages because farmers in hamlets cared better for their 

crops, earned more income from sales and therefore could invest more in fertilizers. Generally, cultivated plots 

near hamlets and villages benefit from greater additions of household refuse and human and animal faeces. 

However, social inequality in access to resources has been found to have an equally large or even larger effect 

on nutrient balances than distance from the village. For instance, positive N, P and K balances were observed 

for Fulani cropland because their large herds supply them with abundant manure. Likewise, nutrient budgets 

ranging from strongly negative to strongly positive were reported for banana-based systems in Tanzania 

depending on access to cattle and cattle management (Roy et al., 2003). Especially in small-holder agriculture, 

site-specific management may also induce large fertility gradients over short distances. 

(Peri-)urban agriculture is characterized by large excesses in nutrients, especially N. This is commonly driven 

by the market-oriented nature of this production system, which allows farmers to invest in external inputs. In 

addition, these systems often rely heavily on the re-use of urban solid waste and waste water. Hence, (peri-) 

urban production systems exemplify another form of large scale fertility transfer, from rural areas to urban 

areas. Food produced by nutrient mining in rural areas is consumed in cities, leading to strong soil enrichment 

of urban soils, especially at urban vegetable production sites (see Box 6.2). 

6.8.5 | What does the future hold? 

Soil nutrient budgets depend on the local socio-economic conditions but also on market prices of inputs and 

on policies. In Western Europe for instance, rising prices of fertilizers and the strengthening of environmental 

policies has led to reductions in N and P inputs into farmland, and this trend is expected to continue. Dwindling 

P resources and climate change may further affect soil nutrient balances, in managed but also in natural 

ecosystems. 


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Box 6.2 | Nutrient balances in urban vegetable production in West African cities 

Based on a two year study of urban gardening sites in Niamey (Niger), it was found that N, P and K 

balances were all positive, with values for high and low input gardens respectively of 1133 and 290 kg ha -1 

for N; 223 and 125 kg ha -1 for P; and 312 and 351 kg ha 

-1 for K. Similar N and P balances were reported for 

urban vegetable gardens in Kano (Nigeria), Bobo Dioulasso (Burkina Faso) and Sikasso (Mali). However, at 

these latter sites, K balances tended to be negative. Overall, urban vegetable production sites appear to be 

major nutrient sinks from which large environmental externalities can be expected. 

Bouwman, Beusen and Billen (2009) evaluated the impact of four future development scenarios on 

nutrient balances for the year 2050. The scenarios, describing contrasting future development in agriculture 

nutrient use under changing climate, are based on the Millennium Ecosystem Assessment. In the most 

pessimistic case, the global N balance may increase by 50 percent in the coming decades. In case of proactive 

policies aiming at closing the nutrient balance, the N balance is expected to remain constant at 150 Tg yr -1 . 

Regarding P, all scenarios predict a future increase in global soil P balance. These global balances hide large 

variations across regions and even across land uses. Unfertilized rangelands are likely to maintain negative 

P balances. Scenarios with a reactive approach to environmental problems portray significant increases in N 

and P balances in Asia, Central and South America and Africa, which can be strongly reduced by a proactive 

approach. For North America, Europe and Oceania, a shift from reactive to proactive environmental policies 

could allow limiting the increase in N and P balances, or even a decrease in the overall nutrient balance. 

Whereas large positive nutrient balances sustained for extended periods of time in industrialized countries 

have resulted in negative environmental externalities, positive nutrient balances should not be viewed as 

necessarily environmentally harmful. Indeed, in many developing regions (e.g. sub-Saharan Africa), positive 

P balances are needed to restore soil fertility potential depleted by long lasting nutrient mining and to boost 

the often very low crop yields. Inputs of N in organic form may also be beneficial as part of a strategy to restore 

the soils’ organic carbon stocks. Possible negative environmental externalities should be weighed against the 

benefits of food security, economic welfare and social well-being. To minimize the negative externalities, the 

best nutrient management approaches should be promoted through judicious policies. 

6.9 | Soil compaction status and trends 

Soil compaction is an important problem affecting productivity of soils across the globe. A hidden problem 

of soils occurring on or below the surface, compaction impairs the function of the subsoil by impeding 

root penetration and water and gaseous exchanges (McGarry and Sharp, 2003). Soil compaction reduces soil 

macroporosity e.g. from an optimum of 6 to 17 percent, and hence reduces pasture and crop yield (Drewry, 

Cameron and Buchan, 2008). 

Soil compaction in most circumstances is a function of soil type (texture, mineralogy, organic matter), 

soil-water content and land management (e.g. tillage practices, traffic, grazing intensity). The problem is not 

limited to crop land but is also prevalent in rangelands and grazing fields, and even in natural non-disturbed 

systems. Soil compaction occurs when compressible soils are subjected to traction e.g. in forest harvesting, 

amenity land use, pipeline installation, land restoration, wildlife trampling (Batey, 2009) or winter grazing 

(Tracy and Zhang, 2008). 

Trampling mechanically disrupts soil aggregates and reduces aggregate stability (Warren et al., 1986) and 

its effect increases with stocking intensity (Willatt and Pullar, 1983). The degree of damage associated with 


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trampling at a particular site depends on soil type (Van Haveren, 1983), soil water content, seasonal climatic 

conditions (Warren et al., 1986), and vegetation type (Wood and Blackburn, 1984). Climate is therefore an 

important determinant of the effects of compaction. Where soil moisture deficits are large, a restriction in 

root depth may have severe effects but the same level of compaction may have a neglible effect where soil 

moisture deficits are small (Batey, 2009). 

Soil compaction effects are long lasting or even permanent (Håkansson and Lipiec, 2000). Especially in 

cultivated land, soil compaction is exacerbated by low soil organic matter content. Intensive use of farm 

machinery including tillage implements such as the mould board, disc ploughs and disc harrows contributes 

to soil compaction, depending on the pattern of load and stress applied and the number of passes. The initial 

condition of the soil also plays a role, including soil moisture, organic matter content, bulk density, particle 

size distribution (including high silt content), and aggregate stability (Materechera, 2008; Horn et al., 2005; 

Imhoff, Da Saliva and Fallow, 2004). Alfisols, a major soil used for crop production in the tropics and covering 

approximately 4 percent of the African land mass, are particularly vulnerable. They are strongly weathered 

and inherently of low organic matter and nutrient status, have a weak structure, and are highly susceptible to 

crusting, compaction and accelerated erosion (Lal, 1987). 

Soil compaction decreases soil physical fertility by impairing storage and supply of water and nutrients, and 

by increasing erosion hazards and the transport of phosphorus and other nutrients out of the farming system. 

Soil compaction can reduce crop yields by as much as 60 percent (Sidhu and Duiker, 2006). The range of yield 

effects is variable, and depends partly on the crop. Cotton was found to be more sensitive to soil compaction 

than were soybeans, corn or Brachiaria brizantha (Busscher, Frederick and Bauer, 2000). Yields of sugarcane 

(Saccharum officinarum L.) were reduced by 40 percent with sub-surface compaction of a clay soil (Jouve and 

Oussible, 1979), while in a clay loam soil wheat yields were reduced by 12 to 23 percent (Oussible, Crookstone 

and Larson, 1992). The compaction effects on yield are greatest when the crop is under stress, such as from 

drought or an excessively wet growing season (Sidhu and Duiker, 2006). Krmenec (2000) observed stand 

count reductions of 20 to 30 percent, plant height decreases of up to 50 percent and yield reductions of about 

19 percent in compacted compared to non-compacted plots. The study of Voorhees, Nelson and Randall (1986) 

illustrates that a one-time compaction event can lead to reduced crop yields up to 12 years later. In another 

study, soil compaction reduced grass yield by up to 20 percent due to N-related stresses (Smith, McTaggart 

and Tsuruta, 1997; Douglas, Campbell and Crawford, 1998). In addition, the creation of waterlogged zones or of 

dry zones caused by shallow rooting can deny plants access to deeper reserves of water (Batey and McKenzie, 

2006). 

Additional consequences include chemical changes, such as the amount of greenhouse gases (nitrous 

oxide and methane) emitted from or taken up in a soil (Hansen, Maehlum and Bakken, 1993; Ruser et al., 1998), 

and reduced root growth and consequently lower crop yields. A study by Gray and Pope (1986) showed also 

that the incidence of Phytophthora root rot in soybeans (Glycine max. L.) was greater with soil compaction. 

Soil compaction increases the abundance of anaerobic microsites and decreases the proportion of coarse 

pores, which may favour emissions of both CH 4 

and N 2 

O (Ball, Scott and Parker, 1999a). Only rarely has soil 

compaction been associated with positive impacts, such as increasing the plant-available water capacity of 

sandy soils (Rasmussen, 1985) or reducing nitrate leaching (Badalıkova and Hruby, 1998) or benefiting soybean 

grown in areas prone to iron deficiency chlorosis in wet years (DeJong-Hughes et al., 2001). 

6.9.1 | Effect of tillage systems on compaction 

While all tillage methods tend to reduce soil bulk density and penetration resistance to the depth of tillage 

(Erbach et al., 1992), equipment used in modern agriculture causes soil compaction of topsoil and subsoil. 

Working the soil to avoid compaction requires timing of tillage in relation to soil water moisture content 

and soil texture (Håkansson and Lipiec, 2000). No-tillage (NT) agriculture is gaining wide acceptance and 


138

is among the top options in the portfolio of technologies to reduce tillage costs, conserve soil and water, 

increase soil organic carbon (SOC) pools, and reduce net CO 2 

emissions, which contribute to global warming 

(Lal et al., 2004). Despite the numerous benefits of NT, there is no consensus yet on its role in alleviating 

soil compaction: some researchers report increased compaction associated with the practice (Bueno et al., 

2006) and others a decrease in compaction (Gregory, Shea and Bakko, 2005). Increasing soil organic matter, 

as practiced in conservation agriculture, reduces soil compactibility (Thomas, Haszler and Blevins, 1996), but 

residue availability remains a key challenge, especially in Africa. 

6.9.2 | What is the extent of deep soil compaction? 

Soil compaction affects mainly topsoils (Balbuena et al., 2000; Flowers and Lal, 1998) but can also affect 

subsoils at depths >30 cm. Most subsoil compaction occurs when the soil is wet and field equipment weights 

exceed 10 tons per axle. The average weight and power of vehicles used on farms has approximately tripled 

since 1966 and maximum wheel loads have risen by a factor of six (Chamen, 2006). While remediation of 

shallow compaction is possible, for example by ripping and subsoiling, correcting soil compaction at depths 

below 45 cm is challenging (Batey, 2009; Berli et al., 2004). Both topsoil and subsoil compaction have been 

acknowledged by the European Union as a serious form of soil degradation, estimated to be responsible 

for degradation of up to 33 million ha in Europe (Akker and Canarache, 2001). Similar compaction problems 

have been reported elsewhere, including in Australia, Azerbaijan, Japan, Russia, China, Ethiopia and New 

Zealand (Hamza and Anderson, 2005). The total amount of compacted soil worldwide has been estimated 

at approximately 68 million ha or around 4 percent of the total land area (Oldeman, 1992; Soane and Van 

Ouwerkerk, 1994). Nearly 33 million ha is located in Europe, where the use of heavy machinery is the main 

cause. Cattle trampling and insufficient cover of the top soil by natural vegetation or crops account for 

compaction of 18 million ha in Africa, and 10 million ha in Asia (Flowers and Lal, 1998; Hamza and Anderson, 2003). 

Agricultural mismanagement (80 percent) and overgrazing (16 percent) are the two major causative factors of 

human induced soil compaction (Oldeman, 1992). 

6.9.3 | Solutions to soil compaction problems 

Soil compaction, like soil chemical characteristics, should be monitored routinely and corrected as part of 

soil management (Batey, 2009). Although soil compaction effects on soil biodiversity and related functions 

and processes depend on several site and soil properties, a threshold of effective bulk density of 1.7 g cm–3 is 

the maximum above which only negative effects are observed (Beylich et al., 2010). Managing soil compaction 

can be achieved through appropriate application of some or all of the following techniques: (a) addition and 

maintenance of adequate amount of soil organic matter to improve and stabilize soil structure (Heuscher, 

Brandt and Jardine, 2005); (b) guiding, confining and minimizing vehicular traffic to the absolutely essential 

by reducing the number and frequency of operations, and performing farm operations only when the soil 

moisture content is below the optimal range for the maximum proctor density (Kroulik et al., 2009); (c) 

mechanical loosening such as deep ripping (Hamza and Anderson, 2005); and (d) selecting a rotation which 

includes crops and pasture plants with strong tap roots able to penetrate and break down compacted soils 

(Hamza and Anderson, 2005). Promoting macrofauna activity can accelerate creation of channels for water 

infiltration and root growth. Arbuscular mycorrhiza can to some extent alleviate the stress of soil compaction. 

This effect has been observed on wheat growth following increased root/shoot ratio of wheat under 

compaction (Miransari et al., 2008). In the long-term, soil compaction can be reduced by natural processes 

that cause the soil to shrink and swell such as wetting and drying (Shiel, Adey and Lodder, 1988), and freezing 

and thawing (Miller, 1980). 

Soil moisture lower than the plastic limit is desirable for cultivation. Traffic should be avoided or restricted when 

condition are otherwise. For farmers, a simple test to avoid soil compaction involves squeezing a small lump of 

soil into a ball and rolling it into a rod about 3 mm in diameter. If a rod can be made easily, the soil is too wet and 

will compact if it is worked or has animals or machinery on it. If the rod is crumbly the water content should allow 

traffic and cultivation without compaction. If a rod will not form at all, the soil could be too dry for tillage in a sandy 

or loamy soil. This test should be run at several points over the full depth of any proposed cultivation. 


139

6.10 | Global soil-water quantity and quality: status, processes and trends 

The world relies on its freshwater for ecosystem health and human well-being and prosperity. Yet only 2.5 

percent of the world’s water is fresh, and of that, 68.7 percent is in the form of ice. Groundwater comprises 30.1 

percent of the freshwater, and just 0.4 percent of the world’s freshwater is in lakes, rivers and the soil. 

6.10.1 | Processes 

Soil water comprises only 0.05 percent of the world’s store of freshwater. However, the upward and 

downward fluxes of water and energy through the soil are massive, and they are strongly linked. The flows are 

upward in the form of water vapour, long-wave radiation and reflected short-wave radiation, and downward 

in the form of liquid water and short-wave radiation (Figure 6.12). The soil-vegetation system is the first 

receiver of the rain and energy that fall on our lands. The soil-vegetation system, which encompasses the 

upper reaches of the groundwater or basement rock to just above the soil-vegetative layer, is the critical zone 

for controlling terrestrial water quantity and quality. 

Rodell et al. (2015) estimate the total annual precipitation onto continents to be 116 500±5 100 km 3 yr -1 

– equivalent to approximately five-times the water stored in the Great Lakes of North America. Sixty percent 

of this (70 600±5 000 km 3 yr -1 ) returns to the atmosphere through evapotranspiration. The remaining 40 

percent (45 900±4 400 km 3 yr -1 ) leaves the continents as runoff, with the greatest proportion either running 

off the surface of the soil or returning to streams via the groundwater flow system after passing through the 

soil. Thus small changes due to human intervention and climate change that alter these fluxes can have very 

large impacts on the store of soil water. 

The quantity, quality and flow of water over and through soil affect the spatial and temporal availability 

and usage of water. The quantity of soil water in a particular layer of soil can be determined by the soil-water 

retention curve, the so-called ‘soil-water characteristic’ (Figure 6.13). This curve describes the relationship 

between the amount of water a particular soil can hold and the energy, or matric potential, required to 

overcome adhesive and cohesive forces to extract water from the soil. Soils of different textures have very 

differing characteristic curves (Figure 6.13) and this affects the movement and storage of water in the landscape. 

© European Space Agency 

Figure 6.12 The flows of water and energy through the soil-vegetation 

horizon 


140

The quality of the soil’s water is determined by the impurities and pollutants present in the soil water, which 

may, or may not be adsorbed to and/or exchanged in some part with the soil’s reactive matrix materials. 

The flow of soil water is determined by the gradient in the matric potential, and the soil’s hydraulic 

conductivity, K (cm day -1 ) (Figure 6.14), which describes the ease with which water flows through the soil 

pore space. The hydraulic conductivity curve is highly non-linear and strongly dependent on the soil’s water 

content, θ, and hence matric potential (Figure 6.14). Soil water flow can vary from very slow in soil with small 

pores, to very fast in soil with large interconnected pores. 

Figure 6.13 The soil-water characteristic curve linking matric potential, to the soil’s volumetric water content. 

Source: Tuller and Or, 2003. 

The soil-water characteristic (Figure 6.14) is an important factor affecting soil microbiology and rhizosphere 

ecology. It controls the stability of the spatial and temporal geometry of the soil pore space, which in turn 

defines the allocation of resources to soil biota, the transport of liquids, gases and solutes to and from roots, 

and the diversity of microbial habitats (Hinsinger et al., 2009). The soil micro-organisms are largely aquatic in 

nature and do not inhabit the air-filled pores. They live instead in the liquid phase of the pores, the thickness of 

which is controlled by the matric potential, which also controls the size and distribution of water-filled pores 

that provide the hydraulic connectivity through soils. 


141

Figure 6.14: The soil’s hydraulic conductivity, K (cm day -1 ) in relation to the matric potential, ψ (MPa). As the matric potential becomes 

more negative the soil’s water content drops (see Figure 6.16) which increases the tortuosity and slows the flow of water. Source: 

Hunter College. 3 

The interactions between the structure and physical, chemical and biological components of the soil control 

the myriad soil functions and processes that are essential for healthy soils, ecosystems and human well-being. 

The soil acts as buffer and filter. Indeed, our soil is the world’s largest water filter. And through this buffering 

and filtering, soil controls the quantity and quality of the world’s liquid freshwater. 

6.10.2 | Quantifying soil moisture 

Soil water varies on multiple time and space scales, driven by climate, weather variability, land cover, 

topography and soil type and structure (Figure 6.15). Measuring variations in soil water is challenging especially 

at large scales where the cost of direct measurement would be very high. Long-term measurement networks 

have historically been limited to a few locations globally (Robock et al., 2000). However, with the recognition 

of soil water as an essential climate variable and the realization that in-situ measurements are necessary for 

the calibration and validation of remote sensing, the number of operational monitoring networks is increasing 

(Dorigo et al., 2011). There are also short-term experimental campaigns with multi-scale soil water sampling 

(Crow et al., 2012). For example, the Soil Climate Analysis Network (SCAN) in the United States provides soil 

water measurements for 174 sites across the United States, with some measurements dating back to 1992. 

New technologies such as the COsmic-ray Soil Moisture Observing System (COSMOS) cosmic-ray neutron 

probes (Zreda et al., 2012) have enabled more efficient and larger measurement footprints of the order of 

several hundreds of square meters. 

3 http://www.geo.hunter.cuny.edu/tbw/soils.veg/lecture.outlines/soils.chap.5/soils_chapter.5.htm 


142

At continental scales, the only practical means of estimating soil water is from satellite sensors or 

simulation models. Satellite-based measurements of soil water are generally based on measuring microwave 

emissions that vary because of the sensitivity of the soil dielectric constant to its wetness. These approaches 

use radiative transfer models to simulate the transfer of radiation emitted from the soil through the vegetation 

canopy and atmosphere to the satellite sensor. However, measurements have generally been restricted to the 

top centimetre of the soil column because of the penetration depth of microwave signals for current sensors 

(> 6 GHz). They are also restricted to sparsely vegetated regions. The recently launched Soil Moisture Ocean 

Salinity (SMOS) (Kerr et al., 2001) and Soil Moisture Active Passive (SMAP) (Entekhabi et al., 2010) satellite 

missions improve on this by using L-band (1-2 Ghz) sensors that have penetration depths of the order 5 cm 

and are less restricted by dense vegetation. Estimates from land surface models have also contributed to 

understanding the variation of soil water at large scales (Sheffield and Wood, 2008). These simulation models 

are driven by observations of precipitation, temperature and other meteorology and simulate the surface 

hydrological cycle with soil water as a prognostic state variable. Recent efforts have developed long-term 

simulations of soil water at regional to global scales (Sheffield and Wood, 2007, 2008; Haddeland et al., 2011), 

although uncertainties exist because of missing process representation in the models and because of errors in 

model structure, parameters and the meteorological forcings. 

6.10.3 | Status and trends 

Understanding variations in soil water is critical for a range of applications including drought risk 

management, agricultural decision making, and understanding and attributing climate change impacts. 

Currently, long-term (multi-decadal) time series of soil water which have been developed from models 

and satellite retrievals are being used to understand variability and long-term changes in soil water 

Figure 6.15 Factors controlling soil water spatial variability and the scales at which they are important. Source: Crow et al., 2010) 


143

(Sheffield and Wood, 2008; Dorigo et al., 2012). Figure 6.16(a) shows the spatial variability of soil water globally 

from model simulations, ranging from high values in the wet tropics and northern boreal forests, to the desert 

regions, such as North Africa, the Middle East, central Asia and Australia. Seasonally, soil water varies with 

changes in precipitation (Figure 6.16 b) with the largest variations in the monsoonal regions of south and 

southeast Asia, west and central Africa and the Amazon. From year-to-year, the El Niño Southern Oscillation 

(ENSO) is the main driver of soil water variability globally (Sheffield and Wood, 2011), often leading to drought 

conditions in the Amazon, south Asia, eastern Australia and southern Africa during El Niño years, and to 

drought in the United States southwest and the Horn of Africa in La Niña years. 

Longer-term changes in soil water are mostly driven by changes in precipitation (Figure 6.16 c and d). Global 

warming may be playing a role in drying soil water in some regions, although this is a subject of debate. Over 

the past 60 years, soil water has been generally wetting over the western hemisphere and drying over the 

eastern hemisphere, mostly in Africa, East Asia and Europe. Trends over the past 20 years (Figure 6.16 e and 

f) indicate intensification of drying in northern China and southeast Australia, and switches from wetting to 

drying across much of North America, and southern South America, in part because of several large-scale and 

lengthy drought events. 

6.10.4 | Hotspots of pressures on soil moisture 

Hotspots of pressures on soil water quantity and quality have emerged around the globe. These result from 

changes in soil water driven by climate change and variability, coupled with human pressures on soil water 

through, for example, agricultural intensification and extensification. We describe three hotspots: the North 

China Plain, the Horn of Africa, and the southwestern United States. 

The North China Plain has seen rapid expansion of agriculture driven by population growth and increasing 

demand for food. This area is relatively dry with around 500 mm yr -1 of precipitation and so irrigation from 

groundwater has become an important feature of agricultural intensification. However, groundwater has 

been used at unsustainable rates, with the result that groundwater levels are dropping by over 1 m per year in 

some parts (Kendy et al., 2003). Furthermore, precipitation has decreased over the past few decades (Figure 

6.16 f). Coupled with intensive irrigation and fertilizer application, this has led to declines in soil water quality 

through salinization and nitrogen leaching (Kendy et al., 2003). 


144

Figure 6.16 (a) Global distribution of average soil moisture depth in the top 1 m of the soil. (b) Seasonal variability in soil moisture 

calculated as the standard deviation of monthly mean soil moisture over the year. (c-d) Global trends (1950-2008) in precipitation 

and 1 m soil moisture. (e-f) As for (c-d) but for 1990-2008. Results for arid regions and permanent ice sheets are not shown. Source: 

Sheffield and Wood, 2007. 


145

Drought has plagued many parts of Africa because of high climate variability from year to year. Severe 

droughts in the 1970s and 1980s led to the deaths of hundreds of thousands of people across the Sahel 

(Sheffield and Wood, 2011). Recent droughts in the Horn of Africa have continued to affect millions of people 

(Ledwith, 2011; UN, 2011), driven by an overall decline in rainfall that is expected to continue and may be linked 

to anthropogenic warming of the Indian Ocean (Funk et al., 2008; Williams et al., 2011). Monitoring soil water 

and its impacts on food security in the Horn of Africa is particularly difficult because of the lack of ground 

measurements. Nonetheless, the use of satellite and modelling technologies has the potential to provide 

drought and famine early warning (Anderson et al., 2012; McNally et al., 2013; Sheffield et al., 2014). 

Soil water in the southwestern United States has been affected over the past two decades by frequent 

severe drought events (2000-2002, 2007, 2009), culminating in a three year drought in California (2011- 

2014) with state-wide impacts on agriculture (Howitt et al., 2014). A shortfall in irrigation water owing to a 

depleted mountain snowpack was partly offset by increasing groundwater pumping. Recent analysis using 

Gravity Recover and Climate Experiment (GRACE) satellites has confirmed the resulting massive losses of 

groundwater since the 1980s from the aquifers underlying California’s agriculturally important Central Valley 

(Famiglietti and Rodell, 2013). McNutt (2014) concludes that “... it is this underground drought we can’t see 

that is enduring, worrisome, and in need of attention”. 


Soil water is vital for the health of terrestrial ecosystems and human well-being. Although only a small 

fraction of the world’s water is stored in the soil, the fluxes of water through the soil are massive. 

On the time-scale of years, the El Niño Southern Oscillation is the prime control on the global variability 

in soil water. At longer time-scales, the global pattern of precipitation is the dominant driver in controlling 

changes in soil water. This pattern may be influenced by climate change. 

Global analysis of the changing patterns of soil water has revealed the emergence of three global hotspots 

in terms of quantity and quality. These are the North China Plan, the Horn of Africa and the southwestern 

United States. There will be great challenges to address in these hotspot regions and in other pockets where 

declining soil water quantity and quality is threatening ecosystem health and human well-being. 

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Stable Base for Agriculture, pp. 209–227. Madison, WI, Soil Science Society of America, Inc. 

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Soil change: 

impacts and responses 


Chencho Norbu (Bhutan), David Robinson (United Kingdom), Miguel Taboada (Argentina) 


Marta Alfaro (Chile), Richard Bardgett (United Kingdom), Sally Bunning (United Kingdom), Jana Compton 

(United States), William Critchley (United Kingdom), Warren Dick (United States), Scott Fendorft (United 

States), Gustavo Ferreira (Uruguay), Tsuyushi Miyazaki (Japan), Carl Obst (Australia), Dani Or (Switzerland), 

Dan Pennock (ITPS/Canada), Matthew Polizzotto (United States), Dan Richter (United States), Marta Rivera- 

Ferre (Spain), Sonia Seneviratne (Switzerland), Pete Smith (United Kingdom), Garrison Sposito (United States), 

Susan Trumbore (United States) and Kazuhiko Watanabe (Japan). 


Dominique Arrouays (ITPS/France), Richard Bardgett (United Kingdom), Marta Camps Arbestain (ITPS/New 

Zealand), Tandra Fraser (Canada), Ciro Gardi (Italy), Neil McKenzie (ITPS/Australia), Luca Montanarella (ITPS/ 

EC), Dan Pennock (ITPS/Canada) and Diana Wall (United States). 

Status of the World’s Soil Resources | Main Report The impact of soil change on ecosystem services 

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7 | The impact of soil change 

on ecosystem services 

7.1 | Introduction 

Soils are now recognized to be in the ‘front line’ of global environmental change and we need to be able to 

predict how they will respond to changing climate, vegetation, erosion and pollution. This requires a better 

understanding of the role of soils in the Earth system to ensure that they continue to provide for humanity 

and the natural world (Schmidt et al., 2011). Although only a thin layer of material at the Earth’s surface, soils 

like many interfaces play a pivotal role in regulating the flow and transfer of mass and energy between the 

atmosphere, biosphere, hydrosphere and lithosphere. Moreover, the structure and organization of soils leaves 

an important imprint on the Earth’s surface in terms of how land is used and how ecosystems develop. Soils 

help regulate the Earth’s physical processes such as water and energy balances, and act as the biogeochemical 

engine at the heart of many of the Earth system cycles and processes on which life depends. Some soil processes 

contribute directly to the delivery of ecosystem goods and services, while other soil processes influence the 

delivery of goods and services. This section examines how soil processes affect soil and ecosystem function 

and the production of goods and services of benefit to humanity. 

Humanity has had an indelible impact on the Earth’s surface, so much so that it has been proposed that the 

planet has entered a new geological epoch, the Anthropocene (Crutzen, 2002). A population of ca. 7 billion 

people that will likely grow to 9.6 billion by 2050 is stressing Earth’s resources. Maintaining the planet in an 

equitable state for human life is perhaps our greatest challenge. Currently, humans have adapted 38 percent 

of the earth’s ice-free land surface to agriculture, crops and pasture (Foley et al., 2011). Agricultural production, 

driven by the need to produce food for a growing population, has had a tremendous impact on our ecosystems 

and resources, especially through the abstraction of water and the leaving of residues. Rockström et al. (2009) 

proposed that we need a ‘safe operating space for humanity with respect to the Earth system’. They argue that 

that there exist biophysical planetary boundaries (or thresholds) which it is inadvisable to cross if we are to 

maintain the needed balance. Vince and Raworth (2012) adapted these concepts to include social goals (1). 

This presentation underlines the fact that we live in a coupled human earth system. The ecosystem services 

analytic approach has been developed in order to bridge the science/policy divide. The approach aims to make 

the concepts clear for all and to set out what needs to be considered in order for humanity to live within 

sustainable boundaries. 


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Soils and soil security are at the heart of this effort. Soil security is defined in McBratney, Field and Koch 

(2014) as “maintaining and improving the world’s soil resources to produce food, fibre and freshwater, to 

contribute to energy and climate sustainability, and to maintain the biodiversity and the overall protection of 

the ecosystem”. Soils perform important ecosystem services (e.g. functions for humanity) including: biomass 

production; storing, filtering and transforming nutrients and water; maintaining a gene pool; providing a 

source of raw material for products such as bricks and tiles; regulating climate and hydrology; and providing 

an archive of cultural heritage. Soils provide ecosystem goods and services directly but some soil processes can 

have an adverse impact on the delivery of ecosystem goods and services. The ability of soils to function can 

be threatened by human activity (on this, see the Soil Thematic Strategy, SEC, 2006). A growing population, 

resource extraction, agricultural production, land use change and climate change all contribute to this threat. 

As population increases, food security is becoming more important in the global agenda. Our historical 

solution to producing more food has been to mechanize, cultivate more land, and increase the application 

of plant nutrients and water. This has led to an almost linear increase in production over time (Pretty, 2008). 

However, the rate of increase is likely to plateau, as has already been seen with wheat in Northern Europe 

and with rice in Korea and China (Cassman, Grassini and Wart, 2010). In addition, agricultural growth comes 

with environmental costs or externalities, which are costs not accounted for in the cost of production. The 

degradation caused can adversely affect everyone, and even the production systems themselves - for instance, 

declines in pollinators can threaten future production (Deguines et al., 2014). 

Figure 7.1 The 11 dimensions of society’s ‘social foundation’ and the nine dimensions of the ‘environmental ceiling’ of the planet. 

Source: Vince and Raworth, 2012. 


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Rockström et al. (2009) suggest we are approaching the limits of the planet’s cultivatable land, while 

the addition of nutrients, especially nitrogen, continues to overload many terrestrial-aquatic systems (Diaz 

and Rosenberg, 2008). At the same time, arable production has seen declines in the carbon content of soils 

- the largest terrestrial carbon reservoir – and these declines are affecting other soil functions, including 

water and nutrient retention (Reynolds et al., 2013). Food production systems will need to change to create 

multifunctional agro-ecosystems capable of maintaining a balance between yields, soil functions and 

biological diversity. Within the field of ecology, this challenge has led to a rigorous debate concerning the loss 

of natural species from agricultural lands - often termed, the ‘land sparing, land sharing’ debate (Green et al., 

2005). This debate has now been integrated within the broad ecosystem services discussion whose central 

ten also focuses on human interaction with ecosystems and their long-term sustainability and continued 

functionality (MA, 2005). 

Conceptually, Foley et al. (2005) proposed that a natural ecosystem provides a range of goods and services 

(2) while on intensively farmed agricultural land, crop production dominates at the expense of all other goods 

and services. They proposed that an ideal situation would be one of balance, with the system producing a 

range of goods and services including food – the ‘sharing’ side of the land sparing/ land sharing debate. 

Organic agriculture has been seen as a model of this sharing or balance. However, organic agriculture has 

so far generally failed to maintain productivity levels in either crop or livestock systems (Pretty, 2008). The 

implication is that organic agriculture does not yet promise balance, because it requires more land and more 

use of natural capital to maintain production levels. Determining if there are viable ‘sharing’ systems should 

continue to be an important research goal but for the moment ‘sparing’ appears to have the upper hand in the 

debate (Phalan et al., 2011). But how do we achieve sustainable intensification? While the viability of sharing 

remains in question, should we focus on a narrow-minded, single service supply management strategy, e.g. 

arable soils for crop production or peat soils for carbon storage? Sustainable intensification research, which 

seeks to find ways of optimizing production while blending in new strategies for multifunctional ecosystem 

service management, is being championed as a way forward (Firbank et al., 2013). 

There is no single solution. Foley’s conceptual diagram (2) highlights the challenges and possible tradeoffs: 

a natural ecosystem delivers a wide range of ecosystem services but scant production; and an intensive 

cropland system delivers royally on production but precious little on ecosystem services. A balanced system 

of cropland with restored ecosystem services would deliver on all services, including production. A recent 

synthesis and analysis of data from the Countryside Survey, a national survey of Great Britain, suggests that 

Foley’s conceptual diagram of intensive cropland (2) is the current situation. Different services reach optimums 

at different points along the productivity gradient, but we cannot have everything (3). The ecosystem service 

indicators alter, often in a non-linear way with the proportion of intensive land use – but with exception of 

production, they all decline with intensification. 3 b and c go on to show that changes in moisture inputs or 

moisture regime, or alteration of soil pH would change the service delivery balance. At no point do we get 

everything, so we will need to choose priorities with our current systems. 

Figure 7.2 Conceptual framework for comparing land use and trade-offs of ecosystem services. Source: Foley et al., 2005. 


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The land sparing, land sharing approach can also been framed in terms of resilience (sharing) and efficiency 

(sparing). Efficient systems by their very nature will prioritise the performance of one function over that of 

others. The degree to which others are affected will depend on whether they perform well under similar 

management or not. Currently, the data in 3 indicate that the choice lies between efficiency and redundancy. 

We can have an efficient carbon storage system, e.g. peat development, which may also perform well as a 

climate thermal buffer because the conditions for peat accumulation require lots of water, but it will not be 

productive for crops in that state, nor will the arable system have high biodiversity as this is inefficient. 

Choices need to be made as to what types of systems we wish to promote. In light of this, the focus of this 

chapter is to assess the global scientific literature and understand how soil change discussed in Chapters 5 

and 6 is likely to impact soil functions and the likely consequences for ecosystem service delivery. Each section 

of this chapter outlines key soil processes involved with the delivery of goods and services and how these are 

changing or - where evidence permits - may change. Each section then reviews how this change impacts soil 

function and affects ecosystem service delivery. Some soil change does not produce an ecosystem service, but 

does impact it; these impacts are considered when assessed as important and adverse. The focus is on the 

local, regional and global scales and follows the general reporting categories of the MA (2005) modified by 

TEEB (2014) to provisioning, regulating and cultural services. Towards the end of the section there is a focus on 

the links with policy, institutions and management. 

7.2 | Soil change and food security 

Keating et al. (2014) provide a useful frame for examining the main roles of soils in food supply through their 

development of the food wedge concept. The food wedge is the triangular area between the level of food 

demand in 2010 and the upper bound of food demand in 2050 (suggested by Keating and Carberry (2010) as a 

wedge equal to approximately 127 x 1015 kcal). The food wedge presented by Keating et al. (2014) assumes that 

food supply and demand were broadly in balance in 2010. Increases in food supply (through, for example, the 

strategies suggested by Foley et al., 2011) would increase the supply to meet the rising demand for food. 

Either the incremental loss of productivity from current agricultural land or the total loss of agricultural 

land due to degradation in the future would cause the lower boundary of the wedge to decrease and hence 

increase the gap between food supply and demand (Figure 7.4). This decrease (or total loss) could occur if the 

services for plant production supplied by the soil decreased due to a significant impairment of one or more 

of the soil functions. Alternatively the restoration of productivity to previously degraded land would increase 

plant production in addition to addressing the yield gap or increases in food delivery. Therefore a key soilfocused 

strategy is to reduce future productivity loss from agricultural soils due to degradation to a minimum 

and to restore productivity to soils that have previously experienced productivity losses. 


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0.0 Response 

1.0 

Water quality 

Soil diversity 

Pollination (B'flies) 

Plant diversity 

Pollination (Bee) 

Freshwater diversity 

Soil C storage 

cSLA 

Cultural 

Figure 7.3 Response curves of mean ecosystem service 

indicators per 1-km 2 across Great Britain. Source: Maskell et 

al., 2013. 

The curves are fitted using generalized additive models to ordination 

axes constrained by; (a) proportion of intensive land (arable and 

improved grassland habitats) within each 1-km square from CS field 

survey data; (b) mean long-term annual average rainfall (1978–2005); 

and (c) mean soil pH from five random sampling locations in each 1-km 

square. All X axes are scaled to the units of each constraining variable 

0.0 Proportion of intensive land in 1km square 

1.0 

0 Response 

1.2 

cSLA 




Butt 


50 Rainfall (mm) 

5000 

0.0 Response 

1.0 


Water quality 


Pollination (B'flies) 



Pollination (Bee) 

cSLA 

Cultural 

2.0 Mean Soil pH 

9.0 


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The restoration of productivity on degraded soils can be complex insofar as soils may have been degraded 

to the point where they cannot readily respond to fertility-improving management techniques. These 

complex interactions among inherent soil properties, management history and the response to inputs is 

well illustrated in the work of Rusinamhodzi et al. (2013) on maize production intensification on smallholder 

farms in Zimbabwe. In this region two major controls of productivity exist – significant differences in yield 

between sandy and clay soils (e.g. inherent soil properties); and pronounced fertility gradients between more 

productive fields close to the homestead and more degraded soils in outfields further from the homestead 

(a management-induced fertility gradient common in many areas of Africa). The sandy soils required longterm 

additions of manure to restore soil functions before the benefit of the mineral fertilizer additions could 

begin to be realized; however even after nine years of substantial organic inputs, the highly degraded sandy 

outfields did not recover their productivity. The authors speculate that the initial soil organic carbon levels 

in the sandy outfields were too low for yields to recover. Moreover at the village scale, the overall amount of 

manure produced is insufficient to apply the required amounts of manure in all fields. 

Figure 7.4 The food wedge and the effect of soil change on the area of the wedge. Source: Keating et al., 2014. 

The relative sizes of the effects of soil change on the food wedge are not drawn to scale. 


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One approach to maintaining soil health is ‘conservation agriculture’, which comprises a range of 

agricultural practices that include reduced tillage and no-till, greater retention of crop residues, and crop 

rotations. However, the lack of organic inputs which constrained productivity in the Zimbabwe example 

above also limits the ability of conservation agriculture to restore fertility in sub-Saharan Africa generally. 

Palm et al. (2014) found that the greatest obstacle to improving soil functions and other ecosystem services in 

Sub-Saharan Africa region is the lack of residues produced due to the low productivity of the soils. The limited 

supply of crop residues also highlights the need to make optimum use of all sources of organic inputs, such as 

animal manure and properly processed human wastes. 

These studies emphasize the inability of mineral fertilizers alone to significantly increase food production 

in regions where the yield gap is greatest. Removing the nutrient limitations through additions of mineral 

fertilizers alone will also exacerbate the range of environmental issues (e.g. N 2 

O emission from N-fertilizer, 

surface and groundwater contamination) in all food-producing regions unless the efficiency of crop use of 

agricultural inputs can be increased. Additionally the fraction of P available as mineable phosphate rock is 

finite. Recent concerns that the world’s supply of phosphorus was being rapidly depleted and that ‘peak 

phosphorus’ was only a few decades away (Cordell and White, 2010) have been dispelled, due to recent 

upward revisions of world phosphate rock reserves and resources (Van Kauwenbergh, 2010). However, the 

world supply of phosphorus is limited, and rising prices and market volatility are inevitable. More efficient use 

of phosphorus is therefore essential. This overall issue is termed the ‘Goldilocks’ problem by Foley et al. (2011) – 

there are many regions with too much or too little fertilizer but few that are ‘just right’. 

A final strategy is to minimize diversion of agricultural soils to production of non-food crops. Recent largescale 

bioenergy production on land previously used for food production has driven a significant land use 

change and represents a major shift of agricultural soils away from food production. Demand for soybean, 

maize and oil palm for biofuel has been a driver of agricultural land conversion in recent years particularly 

in Latin America. Conversion of existing cropland or the development of new cropland for bioethanol and 

biodiesel production competes with food production and carbon returns to the soil (Foley et al., 2011) and thus 

constitutes a threat to soil and food security. Biofuels produced from crops using conventional agricultural 

practices will exacerbate stresses on water supplies, water quality and land use. In any case, biofuels are not 

expected to mitigate the impact of climate change as compared with petroleum (Delucchi, 2011). 

Threats to the food security dimension ‘availability’ are mainly (but not only) caused by soil and land 

degradation and associated water resources (Khan et al., 2009). This is particularly the situation when the 

degradation is irreversible or very hard to reverse. This may, for example, be the case with severe topsoil losses 

caused by wind or water erosion, terrain deformation by gully erosion or mass movement, acidification, 

alkalinization/salinization, soil sealing, or contamination with toxic substances (Scherr, 1999; Palm et al., 

2007; Mullan, 2013). The resulting loss of productivity will reduce yields from a site, leading to reduced returns 

to producers and, in some cases, abandonment of production at the site. Productivity may be restored, but 

economic considerations may limit the adoption of restorative measures. 

The impact of each threat on specific soil functions relevant to crop production has been covered in Chapter 

6 and is summarized in Figure 7.5. The present chapter will focus on the implications for food security of the 

trends in each threat. 

7.2.1 | Soil erosion 

A summary by den Biggelaar et al. (2003) suggests that global mean rates of erosion are between 12 to 15 

tonnes ha -1 yr -1 (Table 7.1). The mid-point of this range yields a soil loss of 0.9 mm yr -1 (see Table 7.1), very similar 

to the mean soil loss of 0.95 mm yr -1 calculated by Montgomery (2007). Overall these rates are substantially 

higher than rates of soil formation, and hence pose a long-term global threat to soils (Montgomery, 2007; see 

also Section 6.1 above). 


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Our understanding of the rates for the three erosion agents (wind, water and tillage) is uneven. Erosion 

rates due to water erosion remain very high (> ca. 20 tonnes ha -1 yr -1 ) in cropland in many agricultural regions 

(Figure 7.2); essentially any cropped area with hilly land and sufficient precipitation is at risk. No reliable global 

estimates for current wind erosion rates exist, and the estimates of the human contribution to current dust 

emissions range from only 8 percent in North Africa to approximately 75 percent in Australia (see also Section 

6.1 above). Tillage erosion primarily results in in-field redistribution of soil, and decreases the productivity of 

soils in convex slope elements and near-upslope field or terrace borders. Global-scale summaries also require 

consideration of the fate of eroded soil – in some regions deposition of eroded soil in river floodplains and 

deltas creates areas of very high and enduring fertility. 

The effect of soil erosion on individual soil properties related to crop production is well documented, but 

the aggregate effect of soil loss on crop yields themselves is less firmly established. The four integrative 

studies summarized in Table 7.1 are based on data sources which range from experimental plot data to reinterpretation 

of GLASOD data. The range of estimates of annual crop loss due to erosion ranges from 0.1 

percent to 0.4 percent, with two studies estimating 0.3 percent yield reduction. 

If the median value of 0.3 percent annual crop loss is valid for the period from 2015 to 2050, a total reduction 

of 10.25 percent could be projected to 2050 (assuming no other changes such as the adoption of additional 

conservation measures by farmers). Foley et al. (2011) cite a value of 1.53 billion ha for cropland globally; the 

loss of 10.25 percent of yield due to erosion would be equivalent to the removal of 150 million ha from crop 

production or 4.5 million ha per year. 

Figure 7.5 Direct impacts of soil threats on specific soil functions of relevance to plant production. 


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Author Database Used Extent Estimates 

den Biggelarr et al. 

(2003) 

Erosion: 179 plot-level 

studies 

Crop yield-erosion: 362 

Global (37 countries) 

Erosion: Average rates 

between 12 – 15 t ha -1 yr -1 

(0.8 to 1.0 mm per year) 1 

Relative annual crop 

yield reduction due to 

erosion: 0.3 percent per 

year (for six major crops) 

Bakker, Govers and 

Rounsevell (2004) 

Erosion-yield: 24 

experimental studies 

Primarily North America 

+ Europe 

Yield reductions of 

approximately 4 percent 

per 10 cm soil loss (= 

0.36 percent per year) 1 

Scherr (2003) 

28 regional studies 

and 54 national or 

sub-national studies 

on soil degradation 

(many GLASOD based, 

primarily soil erosion) 

Global 

Productivity losses 

since WWII of about 

0.3 percent per year for 

cropland and 0.1—0.2 

percent for pasture. 

Crosson (2003) 

Re-analysis of GLASOD 

and Dregne and Chou 

(1992) 

Global Cumulative loss of 5 

percent of productivity 

on 4.7 billion ha 

of cropland and 

permanent pasture 

in 1945 -1 990 period; 

average annual rate of 

loss of 0.1 percent 

Table 7.1 Erosion and crop yield reduction estimates from post-2000 review articles 

1 Calculated using average bulk density of 1.5 tonnes m -3 (den Biggelaar et al., 2003) and average erosion rate of 13.5 tonnes ha -1 yr -1 

(mid-point of den Biggelaar et al., 2003 range) 

The regional differences in crop response to erosion are, however, major. There are great disparities in the 

sensitivity of soils to erosion – soils with growth-limiting sub-soil layers (e.g. shallow soils over bedrock, soils 

with high sodium and/or dense B horizons) are inherently more susceptible to yield reductions due to soil loss 

(Bakker et al., 2007). In a study modelling the impact of erosion in Europe over the next century, Bakker et al. 

(2007) predicted yield reductions on the order of 6 to 12 percent in southern Europe and reductions of only 0 to 

1 percent in much of northern Europe. The overall impact on European food production is, however, relatively 

small as the yields from southern Europe are lower to begin with. In addition, increases in climatic extremes 

associated with human-induced climate change may lead to enhanced levels of wind and water erosion, but 

the impact of these changes will differ greatly among regions. 

Finally, the crop yield/soil erosion relationship may be a less critical reason to reduce soil erosion than 

the off-site impacts of erosion, especially the transport of agricultural inputs such as N and P to waterways 

(Steffen et al., 2015). 


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7.2.2 | Soil sealing 

Soil sealing is most commonly associated with the expansion of urban areas and leads to a permanent, 

non-reversible loss of agricultural land. Yields are eliminated, not just reduced and the soil, if completely 

sealed, becomes effectively non-soil. Urbanization of agricultural land should thus be considered as a threat 

to future food production, not only for the loss of good quality agricultural land but also because of the risk 

of soil pollution through waste disposal and acid deposition from urban air pollution (Chen, 2007; Hubacek 

et al., 2009; Clavero, Villero and Brotons, 2011). Blum and Nortcliff (2013) provide a very rough estimate of 

daily losses of soil due to sealing at the global scale of 250-300 km 2 , and suggest this rate could increase due 

to continuing migration of rural dwellers to urban areas. Thus, new policies that favour sustainable rural 

development, oriented to avoid rural-urban migration as well as to support the return to rural areas of people 

living in the cities, could avoid soil degradation and promote food security. 

7.2.3 | Soil contamination 

Soil contamination reduces food security both by reducing yields of crops due to toxic levels of contaminants 

and by causing the crops that are produced to be unsafe to consume. As summarized in Chapter 6 (Section 

6.3), there are worldwide tens of thousands of known contaminated sites due to local or point-source 

contamination. In regions with a long-standing industrial base, the expansion of contamination is limited, but 

in countries undergoing rapid industrialization or resource development the potential for the further spread 

of contamination is great. The tremendous expansion of industry in China is one example of this: 20 million 

ha of China’s farmland (approximately one fifth of China’s total farmland) is estimated to be contaminated 

by heavy metals, and this may lead to a significant reduction in food availability (see also Section 6.3 above). 

Contamination is also severe due to point sources such as Cs pollution from the Fukushima Dai-ichi nuclear 

power plant and the Chernobyl disaster of 1983. Diffuse soil contamination occurs in many regions (Blum 

and Nortcliff, 2013), but is more commonly linked with concerns about food safety rather than significant 

decreases in crop yields. 

7.2.4 | Acidification 

Acidification of agricultural soils is primarily associated with the net removal of base cations (e.g. product 

removal without replacement with ameliorants such as lime) and the direct addition of acidifying inputs (e.g. 

ammonium-based N fertilizer) to inherently low-pH soils, which have a low capacity to buffer added acidity. It 

is most prevalent on ancient, highly weathered soils. Acidification is a significant regional threat in countries 

such as Australia and Vietnam (see Chapters 10 and 15). Liming is an effective response to control acidity of 

surface horizons, but rates of lime addition lag behind required levels even in developed countries like Australia 

(SOE, 2011) and continuing loss of yield occurs. 

7.2.5 | Salinization 

Salinization in a soil progressively reduces crop yields; beyond a certain crop-specific threshold, growth of 

a given crop may fail entirely. The regional summaries in Chapters 9 to 16 illustrate how difficult it often is to 

separate the causes of salinization: whether the saline soils are naturally occurring (primary salinization) or 

the salinization has been caused by inappropriate management, which is often the case with poorly executed 

irrigation programmes (secondary salinization). Estimates from the 1990s place the land area affected by 

primary salinization at approximately one billion ha, and the area of land with secondary salinization at 77 

million ha (Ghassemi, Jakeman and Nix, 1995). 

Salinization is typically associated with arid and semi-arid areas, and may be exacerbated by climate 

change (see also Section 6.5). An increase in irrigated land is commonly suggested as a means to increase 

food production, but poorly designed and implemented irrigation schemes can readily cause an increase in 


178

salinization. Safe design and operation of irrigation systems requires a high level of managerial expertise. 

Irrigation expansion can contribute to increases in food production but great care is needed in planning and 

design to avoid negative effects such as salinization. 

7.2.6 | Compaction 

Compaction impairs soil functions by impeding root penetration and limiting water and gas exchange. In 

soils where it occurs, it can reduce crop yields but it rarely eliminates plant growth entirely. The susceptibility 

of different crops to compaction differs greatly (see also Section 6.9 above). Good soil management requires 

care in minimizing soil compaction and the adoption of management practices which alleviate existing 

compaction. The effect of soil compaction on output and hence on food security is, however, difficult to 

assess, especially in tropical areas (Lal, 2003). 

7.2.7 | Nutrient imbalance 

The problems associated with under-supply of nutrients in regions such as Sub-Saharan Africa will be 

discussed in Chapter 8 in the context of closing the yield gap. Foley et al. (2011) and Steffen et al. (2015) clearly 

indicates the regions where over-supply of nutrients is occurring: mid-west United States, western Europe, 

northern India, and the coastal areas of China. Foley et al. (2011) emphasize the need to address the economic 

and environmental issues in nutrient over-supply by increasing the efficiency of nutrient uptake by plants. This, 

coupled with reductions in transport of nutrients to waterways by minimizing erosion, would substantially 

reduce eutrophication. It would also allow the redistribution of N and P to areas of nutrient-poor soils without 

exceeding the planetary boundaries for the elements (Steffen et al., 2015). 

7.2.8 | Changes to soil organic carbon and soil biodiversity 

Soil organic carbon (SOC) and soil biodiversity are commonly linked to three dimensions of food security: 

increases in food availability, restoration of productivity in degraded soils, and the resilience of food production 

systems. 

Soil C is not itself a direct control on food production but is a proxy for soil organic matter (SOM), which 

is one of the key attributes associated with many soil functions. Soil microbial C is normally included in 

aggregate measures of SOC, and soil microbes are a component of the soil organic matter; hence in terms 

of mass, SOC/SOM and soil microorganisms are directly related. The focus on SOC, rather than SOM, occurs 

because of the ease of measurement of C as a proxy for SOM, and because of the direct connection between 

SOC and atmospheric C. 

The roles of SOC and soil biodiversity in increasing food availability are also inextricably bound together. 

Increases in SOC and in soil biodiversity are believed to be beneficial for crop production, and decreases in both 

are equally believed to be deleterious for crops; however providing evidence for these qualitative statements 

and establishing predictive relationships has been difficult (Naeem et al., 2009; Bommarco, Kleijn and Potts, 

2013; Palm et al., 2014). 

The more readily understood relationship between soil C storage and atmospheric C levels has driven much 

of the work in the past 15 years on soil carbon dynamics, but the secondary benefit of increasing SOC levels 

for crop production is commonly cited, if rarely quantified. Efforts to determine a threshold SOC value for 

maximum crop production in temperate soils have not been successful as it depends on management and on 

other factors such as soil limitations and precipitation (Loveland and Webb, 2003). Lal (2006) estimates yield 

gains associated with a 1 Mg ha -1 gain in SOC in the tropics and sub-tropics ranging from 20-70 kg ha -1 yr -1 for 

wheat to 30-300 kg ha -1 yr -1 for maize. However, the study acknowledges that the data are meagre and that 

functional relationships between SOC pool and crop yield are not available, especially for degraded soils in the 

tropics and subtropics. 


179

Research in tropical and semi-tropical lands has established that inputs of organic material through the 

return of residues and manure to the soil are essential for fertility restoration in degraded soils, but that low 

residue production and competing uses for residues and manure limit the adoption of these SOC-aggrading 

approaches (e.g. Lal, 2006; Rusinamhodzi et al., 2013; Palm et al. 2014). Sustainable soil management that 

increases SOM levels will assist in maintaining productivity, but the specific measures taken to increase SOM 

must be locally developed. 

Establishing a direct, quantitative link between soil biodiversity and increasing food production is even more 

elusive. Sylvain and Wall (2011) observe that “the total invertebrates found in a soil will interact to provide many 

services and participate in several ecosystem functions, but it is unlikely that a single species will influence all 

services and functions that influence plant growth or composition at the same time or in the same manner”. 

Biodiversity beyond the soil plays an important role in regulating services such as biological pest control and 

crop pollination (Bommarco, Kleijn and Potts, 2013), and public concerns about the effects of pesticides on key 

species continues to grow. 

A final role for SOC enhancement and maintenance of soil biodiversity is to increase the resilience of the soil 

for food production, especially its ability to withstand disruption due to human-induced climate change. SOC 

buffers the impact of climate extremes on soils and crops by: (i) regulating water supply by reducing runoff 

and increasing soil-water holding capacity; (ii) reducing erosion through runoff reductions and improved 

aggregation; and (iii) providing sites for nutrient retention and release (Loveland and Webb, 2003; Lal, 2006). 

The combined role of soil organic matter and biodiversity in nutrient cycling ensures a continuing supply of 

nutrients for crop growth. It is difficult to quantify this relationship, especially in the light of the uncertainties 

associated with human-induced climate change, but the existing qualitative understanding is sufficient to 

establish the importance of SOC and biodiversity in sustainable soil management. 

Summary 

The importance of soil degradation and soil rehabilitation are highlighted in principles eight and nine of the 

proposed World Soil Charter: 

Soil degradation inherently reduces or eliminates soil functions and their ability to support ecosystem 

services essential for human well-being. Minimizing or eliminating significant soil degradation is essential 

to maintain the services provided by all soils and is substantially more cost-effective than rehabilitating 

soils after degradation has occurred. 

Soils that have experienced degradation can, in some cases, have their core functions and their contributions 

to ecosystem services restored through the application of appropriate rehabilitation techniques. This 

increases the area available for the provision of services without necessitating land use conversion. 

Our ability to predict the effect of soil degradation on food security is very limited for two main reasons. First, 

there is a lack of up-to-date knowledge both on the area affected by degradation and on the linkages between 

degradation and soil functions (and ultimately plant production). The research community continues to cite 

research summaries on the effects of soil degradation on crop yields from the 1990s based on data gathered in 

the 1980s. Yet crop production in many regions has undergone profound change since the 1980s – for example, 

the widespread adoption of conservation tillage in many regions occurred during the 1990s and 2000s. There 

is a pressing need for meta-analyses on all of the soil threats discussed here. This in-depth review of existing 

work needs to be complemented by new research to address major information gaps, and in particular to 

prove more conclusively the functional relationships between soil attributes and plant production. 


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The second limitation to predictions is that farmers are not simply passive observers of inexorable 

degradation processes – farmers in all regions, including the tropics, are willing to invest in the future to 

protect soils and the essential services that they provide (Stocking, 2003). For example, the adoption of 

conservation tillage in heavily mechanized systems such as those in North America has substantially lowered 

erosion rates (Montgomery, 2007). The general applicability of conservation tillage in other regions may be 

limited (Palm et al., 2014) but the principle that farmers are active participants in soil change is essential to 

recognize and encourage. 

7.3 | Soil change and climate regulation 

Soils play a fundamental role in the maintenance of a climate favourable to life. A range of soil processes 

helps regulate climate, including the thermal and moisture balance, greenhouse gases (H 2 

O, CO 2 

, CH 4 

and 

N 2 

O) and particulates in the atmosphere. Soils can also adversely impact the maintenance of air quality. 

7.3.1 | Soil carbon 

Although it is hard to estimate quantities, it is certain that soils contain vast reserves of carbon. Recent 

estimates range between 1200 and 3000 Pg C depending on the depth to which estimates extend, and on the 

way in which wetland soils are counted (Hiederer and Köchy, 2012). Roughly 1670 Pg of C is stored in peatlands 

and permafrost soils in high northern latitudes (Tarnocai et al., 2008). Hence soil organic matter is a large 

pool. Consequently, only small changes in soil C storage can have a large effect on atmospheric CO 2 

. Soils also 

contain approx. 950 PgC in the form of pedogenic carbonates to 2 m depth (Batjes, 1996). 

Carbon respired from soils and derived from decomposition of organic matter in soils approximately 

balances annual net primary production of carbon by biomass. Carbon dioxide derived from plant roots and 

their symbionts below ground adds to the total flux of CO 2 

from soils to the atmosphere, which in total is ~10 

times larger than the current release of CO 2 

to the atmosphere by fossil fuel burning (Schimel, 1995). Hence 

relatively small changes in the cycling of soil C can lead to large changes in atmospheric CO 2 

. 

Management that changes C inputs or tillage that alters the stability of soil organic matter through changes 

in soil aeration or structure measurably alter soil C storage. Historically, the expansion of agriculture has led to 

losses of soil C to the atmosphere, estimated globally to be of order 40-90 PgC, some of which has remained 

in the atmosphere (Smith, 2004, 2012). 

In terms of climate change, most projections suggest soil carbon changes driven by future climate change 

will range from small losses to moderate gains, but these global trends show considerable regional variation 

(Smith, 2012). The response of soil C in future will be determined by two factors: (i) the impacts of increased 

temperature and altered soil moisture on decomposition rates; and (ii) the balance between increases in C 

losses resulting from accelerated decomposition and predicted C gains through enhanced productivity under 

elevated CO 2 

and nutrient deposition (Smith, 2012). 

Soil organic matter (SOM) is considered dynamic and has importance beyond its climate role. Plant residues 

added to soils provide energy for a cascade of heterotrophic organisms. A key outcome of organic matter (OM) 

breakdown is the release of essential nutrients into the soil. If the breakdown of OM exceeds the supply of 

O 2 

, e.g. under high moisture conditions, the degradation of OM using other electron acceptors drives the 

production and consumption of other important greenhouse gases such as methane and nitrous oxide. The 

degradation of OM also indirectly affects greenhouse gases like troposphere ozone by altering the emission of 

reactive trace gases. In addition to climate effects through regulation of greenhouse gases, SOM determines 

properties such as nutrient retention, water retention, and the structure and size of the microbial community 

in soils. 


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SOM feedbacks to climate change (Figure 7.6) include direct responses such as: (i) the alteration of microbial 

activity with temperature (Conant et al., 2011); and (ii) moisture-related changes in the supply of O 2 

relative to 

other electron acceptors that reflect precipitation change. In this context, probably the biggest concern is the 

thawing of large stores of C in permafrost at high northern latitudes, which will make organic C that has been 

frozen for millennia available for decomposition. This response is predicted to create a significant positive 

feedback to climate change (Schuur et al., 2008). 

Another direct response of soil organic carbon pools is predicted in response to elevated CO 2 

and its effect 

on ecosystem productivity. Free Air Carbon Dioxide Enrichment (FACE) studies have shown that belowground 

productivity can be strongly affected, with cascading and mixed consequences for SOM storage. However 

indirect effects are such as altered stabilization of older C associated with the increased inputs of fresh 

plant inputs (‘priming’) add uncertainty to the prediction of future soil C responses. As with CO 2 

fertilization, 

increased deposition of reactive N associated with regional air pollution affects production, quality and spatial 

distribution of plant inputs (e.g. above- versus belowground) and can alter the decomposition rates through 

changes in the soil microbial community (Berg and Matzner, 1997). Hence the net effects on soil C storage are 

difficult to predict, though the combined effects of climate change and fertilization are expected to result in 

net losses of soil C overall from temperate forest soils (Hopkins, Torn and Trumbore, 2012) 

Many of the processes affecting SOM over the past century have been dominated by human management 

of vegetation, which in turn affects the inputs and status of SOM. Changes in vegetation cover, including 

those occurring in response to climate as well as to land use or management, influence soil organic matter by 

altering the rates, quality and location of plant litter inputs to soils. In turn, litter inputs influence the amount 

and composition of the decomposer organisms, including soil fauna, as well as the soil microbial community. 

Studies of a number of vegetation transitions – for example the replacement of forests with agriculture or 

pasture – have shown that these transitions have led to a loss of soil C to the atmosphere. However, the 

trajectory of vegetation change in response to climate, and the consequences for atmospheric CO 2 

, are not 

well known, as soils will in turn determine what kind of vegetation will take over. For example, C from thawing 

permafrost soils may eventually be sequestered in the biomass of forests that can grow in the warmer climate. 

Evidently the time lags required for these transitions are an important part of understanding the net effect of 

soil C on the carbon cycle. 

In addition to direct effects of changes in plant litter addition to soils, management or vegetation change 

also alters the chemical and physical framework of soil and thereby the organisms inhabiting it. For example, 

ploughing can break up soil aggregates and make organic matter that was previously protected available 

to decomposers. Changes in evapotranspiration can change local and regional water resources. Addition of 

fertilizers increases plant productivity but also alters soil microbial communities and can stimulate production 

of reactive N gases and N 2 

O. 

Large-scale soil erosion is thought to slow decomposition of buried, eroded organic matter, while growth of 

vegetation on the remaining soil will tend to increase soil C storage. However, these effects have been shown 

to be relatively small (Van Oost et al., 2007). By removing topsoil that is generally high in organic matter, 

erosion can have profound effects on physical and chemical soil properties such as water retention and cation 

exchange capacity. 

Global increases in carbon stocks have a large, cost-competitive potential for climate change mitigation 

(Smith et al., 2008). Mechanisms include reduced soil disturbance, improved rotations and residue/organic 

input management, and restoration of degraded soils. Nevertheless, limitations on soil C sequestration 

include time limitation, non-permanence, displacement and difficulties in verification (Smith, 2012). Despite 

these limitations, soil C sequestration can be useful to meet short- to medium-term targets. In addition, soil C 

sequestration confers a number of co-benefits on soils. It is thus a viable option for reducing the atmospheric 


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Figure 7.6 Some soil-related feedbacks to global climate change to illustrate the complexity and potential number of response 

pathways. Source: Heimann and Reichstein, 2008. 

CO 2 

concentration in the shorter term, buying time to develop longer term emission reduction solutions 

across all sectors of the global economy (Smith, 2012). 

Just as reductions in soil C stocks are associated with negative consequences for soil function, increased 

soil carbon stocks are associated with increased soil fertility, workability, water holding capacity, reductions in 

greenhouse gas emissions and reduced erosion risk (Lal, 2004). Increasing soil carbon stocks can thus reduce 

the vulnerability of managed soils to future global warming (Smith and Olesen, 2010). Management practices 

effective in increasing SOC stocks include: (i) improved plant productivity through nutrient management, 

rotations and improved farming practices; (ii) reduced or conservation tillage and residue management; (iii) 

more effective use of organic amendments; (iv) land use change, for example from crops to grass or trees; (v) 

set-aside; (vi) agroforestry; (vii) optimizing livestock densities; and (viii) planting legumes or improving the 

crop mix (Smith et al., 2008). While these measures have the technical potential to increase SOC stocks by 

about 1 – 1.3 Pg C yr -1 (Smith et al., 2007a, 2008), they are dependent on economics: the economic potential for 

SOC sequestration was estimated to be 0.4, 0.6 and 0.7 Pg C yr -1 at carbon prices of up to US$20, $50 and $100 

per tonnes CO 2 

-eq. yr–1, respectively (Smith et al., 2008). In addition, the size of the potential sequestration 

is relatively small in comparison to the threats: only a small loss of C from permafrost or peatlands could 

offset this potential sequestration (Joosten et al., 2014). However, an increase in SOC through improved 

management is expected to also reduce vulnerability of the soils to future SOC loss under global warming. 

As such, soil carbon sequestration can, in many respects, be regarded as a ‘win-win’ and a ‘no regrets’ option 

(Smith et al., 2007b). 

7.3.2 | Nitrous oxide emissions 

Soils emit nitrous oxide (N 2 

O), a greenhouse gas that is around 300 times more potent for radiative forcing 

(climate warming) over 100 years than CO 2 

. Of the approximately 16 Tg N 2 

O-N yr–1 emitted globally in the 

1990s, between 40 and 50 percent was a result of human activities (Reay et al., 2012). The main sources were 


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agriculture, industry, biomass burning and indirect emissions from reactive nitrogen, such as leaching, runoff 

and atmospheric deposition (Reay et al., 2012). Of these sources, agricultural soils are the dominant source, 

contributing over 80 percent of global anthropogenic N 2 

O emissions during the 1990s (Smith et al., 2007a). 

N 2 

O emissions from agricultural soils have increased from just under 4 Tg N 2 

O-N yr–1 in 1990, to over 4 Tg 

N 2 

O-N yr–1 in 2010. Emissions are projected to increase to over 5 Tg N 2 

O-N yr–1 by 2030 (Reay et al., 2012). 

Nitrous oxide is emitted from soils through two processes, nitrification and denitrification. Any mineral 

N available in the soil is subject to loss through one of these processes. The processes depend on soil 

environmental conditions such as the availability of mineral N, soil temperature and soil water content, 

soil pH, organic matter content and soil type. Nitrification tends to be favoured under aerobic conditions 

and denitrification under anaerobic conditions (Galloway et al., 2003). Subject to mineral N being available, 

any soil can emit N 2 

O through mineralisation of soil organic matter. However, the majority of emissions are 

driven by sources of N added to the soil as fertiliser, either as synthetic fertilizer, or as organic amendments 

(e.g. manures, slurries, composts). So close is the relationship between N addition and emission, that N 2 

O 

emissions are often calculated as a direct function of N added to the soil (Reay et al., 2012). Emissions of N 2 

O 

from agricultural soils driven by addition of synthetic fertilizers have increased from 67 MtCO 2 

-eq. yr -1 in 1961, 

to 683 MtCO 2 

-eq. yr -1 in 2010 (Tubiello et al., 2013). 

Given the close association between N inputs and N 2 

O emissions, soil management strategies to reduce 

N 2 

O emissions, and thereby improve this aspect of their climate regulation function, are mostly centred on 

removing surplus N in the soil. This is mainly accomplished by improving N-use efficiency to reduce the N 

surplus, either by reducing inputs or by better matching applications (timing and amount) to plant demand 

(Snyder et al., 2014). In a recent review, Snyder et al. (2014) noted that soil N 2 

O emissions can be reduced by 

selecting the right source, rate, time and place of N application and that new technologies and greater farmer/ 

adviser skills can improve N input management. They estimate that crop N recovery could be increased by >20 

percent, reducing risks of N 2 

O emissions by >20–30 percent (Snyder et al., 2014). 

Beyond these technical measures, N 2 

O emissions could also be reduced through demand-side management, 

for example through reduced food waste. Another demand-side measure could be to encourage dietary 

change away from less efficiently produced food products such as meat and other livestock products, or foods 

with very high energy inputs, such as heated glasshouses during winter (Reay et al., 2012). 

In summary, managed soils can play a key role in climate regulation via N 2 

O emissions, and a number of 

options exist to improve the soil’s delivery of its climate regulation service both by enhanced N management 

and by wider systemic changes in agriculture (Flynn and Smith, 2010; Reay et al., 2012; Snyder et al., 2014). 

7.3.3 | Methane emissions 

Methane (CH 4 

) is a greenhouse gas that is around 20–35 times more potent for radiative forcing (climate 

warming) over 100 years than CO 2 

. Soils often emit methane through methanogenesis when decomposition 

of organic matter occurs in anaerobic soil layers. Methane is also oxidised by methanotrophy in aerobic 

layers, so the emission is a balance between methanogenesis and methanotrophy (Le Mer and Roger, 2001). 

About 30 percent of total global CH 4 

emissions are natural (including the natural wetland flux), and around 

70 percent anthropogenic (Le Mer and Roger, 2001). Given that methanogenesis occurs under anaerobic 

conditions, waterlogged soils, particularly wetlands, peatlands and rice paddies, are the largest source of 

methane emissions (Le Mer and Roger, 2001). Since much of the methane flux from wetland and peatland 


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soils occurs on largely unmanaged areas, the emissions are not considered anthropogenic, so are not routinely 

included in greenhouse gas inventories. This means that the quantification of soil methane emissions over 

time from peatlands and wetlands is not as well documented as for N 2 

O. Nonetheless, some global estimates 

of CH 4 

emissions from wetlands do exist: in 1998, total global emissions of CH 4 

from wetlands were estimated 

to be 145 Tg yr -1 , of which 92 Tg yr -1 came from natural wetlands and 53 Tg yr -1 from rice paddies (Cao, Gregson 

and Marshall, 1998), with some estimates a little higher (Le Mer and Roger, 2001). Emissions from rice paddies, 

however, are included in inventories: CH 4 

emissions from rice paddies were estimated to have increased from 

366 MtCO 2 

-eq. yr -1 in 1961 to 499 MtCO 2 

-eq. yr -1 in 2010 (Tubiello et al., 2013). 

By contrast, aerobic soils tend to act as sinks for CH 4 

, thereby having a positive impact on climate regulation. 

Temperate and tropical aerobic soils that are exposed to atmospheric concentrations of CH 4 

usually exhibit 

low levels of atmospheric CH 4 

oxidation but, since they cover large areas, they are estimated to consume ~10 

percent of the atmospheric CH 4 

(Le Mer and Roger, 2001). Forest soils are the strongest CH 4 

sink, followed by 

grasslands, with the sink capacity of cultivated land much lower than that of undisturbed soils (Steudler et al., 

1996; Priemé et al., 1997). Atmospheric CH 4 

oxidation also occurs in extreme environments such as deserts and 

glaciers, in the floodwater of submerged soils and in river waters (Le Mer and Roger, 2001). Potter, Davidson 

and Verchot (1996) estimated global soil CH 4 

consumption to be 17–23 Tg yr−1. 

Soil management strategies to reduce CH 4 

emissions or enhance CH 4 

uptake can improve this aspect of the 

soil’s climate regulation function. Enhancing uptake in managed soils is difficult, so most mitigation options 

occur for CH 4 

emission reduction, and since wetlands or often unmanaged, most mitigation options have been 

developed for rice paddies. These include draining the wetland rice once or several times during the growing 

season, selection of rice cultivars with low exudation rates, off-rice season water management, fertilizer 

management and the timing and composting of organic residue additions (Smith et al., 2008). For managed 

peatlands and wetlands (e.g. those used for forestry or agriculture), methane emissions can be reduced 

by fertilizer, water and tillage management (Le Mer and Roger, 2001). Rewetting of drained or cultivated 

peatlands to restore wetland function and maintain carbon stocks is likely to increase CH 4 

emissions, but the 

overall impact on climate will vary between systems and depending on the time horizon considered (Joosten 

et al., 2014). 

7.3.4 | Heat and moisture transfer 

Soils play an essential role in storage of water. Soil moisture strongly affects water, energy and carbon 

exchanges, leading to major forcings and feedbacks within the climate system (Seneviratne et al., 2010). Soil 

moisture generally refers to the amount of water stored in the unsaturated soil zone. The most important 

soil moisture storage is that affecting plant transpiration, e.g. the water available within the root zone. Land 

evapotranspiration is an essential component of the continental water cycle, since it returns as much as 60 

percent of precipitated land water back to the atmosphere (e.g. Dirmeyer et al., 2006; Oki and Kanae, 2006; 

van der Ent et al., 2010). Soil moisture is the main water source for this process, through plant transpiration 

and bare soil evaporation. Plant transpiration contributes about 60 percent of all land evapotranspiration 

(Schlesinger and Jasechko, 2014). 

Evapotranspiration is itself a function of soil moisture (Koster et al., 2004; Seneviratne et al., 2010). This 

dependency is conceptually illustrated in Figure 7.7, which builds upon the classical Budyko framework (Budyko, 

1956, 1974). It shows that three main soil moisture regimes can be distinguished: (i) a wet soil moisture regime 

in which evapotranspiration is solely limited by the availability of energy; (ii) a transitional soil moisture regime 

in which evapotranspiration is strongly sensitive to the availability of soil moisture; and (iii) a dry soil moisture 

regime in which soil moisture is at or below the wilting point and for which evapotranspiration is negligible. 


185

The geographical distribution of these soil moisture and evapotranspiration regimes can be estimated 

with various methods, as discussed in Seneviratne et al. (2010). As an illustration, Figure 7.8 displays the 

correlation of annual mean evapotranspiration with radiation and precipitation in an observation-driven 

land surface model using a two-dimensional colour map. This analysis illustrates the existence of distinct 

evapotranspiration regimes, with most regions clearly displaying either the characteristics of a soil moistureor 

energy-limited evapotranspiration regime. 

One should note that the relationship displayed in Figure 7.7 is qualitative, and is affected (both in space and 

time) by variations in soil parameters, land cover characteristics, and other factors (e.g. Teuling et al., 2010; 

Koster and Mahanama, 2012; Guillod et al., 2013). 

The water and energy balances of land are tightly connected through the process of evapotranspiration. It 

follows that the soil moisture effects on evapotranspiration (illustrated in Figure 7.8) are also highly relevant 

for land energy exchanges at the land surface. This link makes soil moisture a strong control of temperature 

variability and temperature extremes on land (e.g. Seneviratne et al., 2006; Fischer et al., 2007; Vautard et 

al., 2007; Mueller and Seneviratne, 2012). Modelling estimates suggest that soil moisture feedbacks affect 

about 60 percent of temperature variability in the present Mediterranean climate in summer (Seneviratne et 

al., 2006) and that they induced additional temperature anomalies of the order of 2°C in Central Europe during 

the 2003 European summer heat wave (Fischer et al., 2007). Observation-based analyses also confirm the 

existence of strong correlations between the occurrence of hot extremes in regional hottest months and prior 

precipitation deficits in regions with soil moisture-limited evapotranspiration regimes (Hirschi et al., 2011; 

Quesada et al., 2012; Mueller and Seneviratne, 2012). The example of the European summer heat wave shows, 

moreover, that these feedbacks can be relevant in extreme years even in regions like Central Europe which 

have a dominant energy-limited evapotranspiration regime under the present climate. 

For present climate conditions, the relationship between soil moisture deficits and hot extremes implies 

that information on soil moisture deficits could be used for improved forecasting of temperature mean and 

Dry 

Transitional 

Wet 

EF max 

Soil Moisture Limited Energy limited 

EF=λE/R n 

0 

θ WILT 

θ CRIT 

Soil moisture content 

Figure 7.7 Definition of soil moisture regimes and corresponding evapotranspiration regimes. Source: Seneviratne et al., 2010. 

EF denotes the evaporative fraction, and EFmax its maximal value. 


186

extremes several weeks in advance (e.g. Koster et al., 2010a; Mueller and Seneviratne, 2010). Such early soil 

moisture information can be either provided by an offline land surface model driven with observation-based 

forcing (e.g. Dirmeyer et al., 2006), by remote sensing products (e.g. Wagner et al., 2007; De Jeu et al., 2008), 

or by the assimilation of remote sensing products in land surface models (e.g. Reichle, 2008). However, the 

scarcity of precipitation and soil moisture observations still limits the derivation of reliable soil moisture 

estimates and the evaluation of satellite approaches on most continents (e.g. Koster et al., 2010b; Dorigo et 

al., 2013). 

Figure 7.8 Estimation of evapotranspiration drivers (moisture and radiation) based on observation-driven land surface model 

simulation. Source: Seneviratne et al., 2010. 

The figure displays yearly correlations of evapotranspiration with global radiation Rg and precipitation P in simulations from the 2nd 

phase of the Global Soil Wetness Project (GSWP, Dirmeyer et al., 2006) using a two-dimensional color map, based on Teuling et al. 

2009, redrawn for the whole globe. (Seneviratne et al., 2010) 

Climate models project that several regions will be affected by more frequent drought conditions in the 

future as a consequence of enhanced greenhouse gas concentrations (e.g. Wang, 2005; Sheffield and Wood, 

2007; Seneviratne et al., 2012). This implies shifts in climate and soil moisture regimes, with important impacts 

on temperature projections (e.g. Seneviratne et al., 2006; Dirmeyer et al., 2012), in particular for temperature 

extremes (Seneviratne et al., 2013). 

Another feedback of soil moisture on climate is the possible impact of droughts on plant carbon uptake 

and a resulting decreased sink for CO 2 

emissions (Ciais et al., 2005; Friedlingstein et al., 2006; Sitch et al., 

2008; Reichstein et al., 2013). One particularly important region for this feedback is the Amazon rainforest, 

which is projected in some models to dry substantially (e.g. Mahli et al., 2008). However, these projections 

are associated with high uncertainty in current climate models (Orlowsky and Seneviratne, 2013), and the 

resulting effects on carbon uptake could also be affected by the representation of plant physiology in the land 

surface schemes (Huntingford et al., 2012). 

Finally, the combined effects of soil moisture on near-surface humidity and temperature are also relevant 

for boundary layer development and precipitation occurrence (e.g. Betts, 2004; Koster et al., 2004; Taylor et 

al., 2012). More details on these feedbacks are provided in Sections 7.5 and 7.6 below. 


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7.4 | Air quality regulation 

According to the World Health Organisation (http://www.who.int/topics/air_pollution/en/), air pollution 

is “contamination of the indoor or outdoor environment by any chemical, physical or biological agent that 

modifies the natural characteristics of the atmosphere”. The status of air pollution is often referred to as air 

quality (Monks et al., 2009). Air quality affects human health through exposure to toxic inorganic compounds 

(e.g. HBr, elemental Hg vapour), toxic organic compounds (e.g. organic pesticides), and particulate matter 

(PM). Air quality also affects the climate system through changes in greenhouse gas concentrations (CO 2 

, CH 4 

, 

N 2 

O) – as discussed in Section 7.3 − and through aerosols (e.g. mineral particles, black carbon or ‘BC’). After 

deposition of atmospheric pollutants (e.g. N and S compounds, or compounds containing trace elements) on 

land or water, acidification, eutrophication, and contamination might occur (see Section 4.4), which can have 

harmful effects on ecosystem function and structure, particularly where deposition exceeds the ‘critical load’ 

that a particular soil can buffer (Nilsson and Grennfelt, 1988). Specific compounds in the atmosphere, such as 

ammonia (NH 3 

), can result in a host of environmental problems (e.g. impacts on human health, odour, climate 

change, soil acidification, eutrophication, biodiversity). The magnitude of the problems would depend on 

interactions with other compounds (Aneja, Schlesinger and Erisman, 2009). 

7.4.2 | Ammonia emissions 

Agriculture accounts for 80–99 percent of all NH 3 

emissions (FAO, 2014). In Europe, agriculture accounts 

for 94 percent (EEA, 2012). These emissions mainly come from animal manure and fertiliser application (Olivier 

et al., 1998). In the United States, NH 3 

reductions are voluntary and there are neither federal nor national 

regulations controlling its emission (Aneja, Schlesinger and Erisman, 2009; Greaver et al., 2012). In Europe, 

however, NH 3 

emissions have been an important policy issue (van der Hoek, 1998) and regulation has led 

to an overall reduction in NH 3 

emissions. Between 1990 and 2010, NH 3 

emissions decreased in the EU-27 by 

28 percent (EEA, 2012), with especially large reductions in Poland, the Netherlands and Germany. Ammonia 

emission reductions have been associated with a reduction in the number of livestock (especially cattle), 

improvement of manure management, and the lower input of nitrogenous fertilisers to soils (EEA, 2011, 2012). 

The effectiveness of manure injection to decrease emissions is under debate, as a result of its effect on pollutant 

swapping, as there may be a reduction in NH 3 

but an increase in N 2 

O emissions and/or NO 3 

leaching (Erisman 

et al., 2008). A better understanding is needed on the contribution of NH 3 

as a precursor of PM concentrations, 

both emissions of primary PM 10 (particulate matter with a size < 10 µm) and secondary formation of PM 2.5 

(particulate matter with a size < 2.5 µm) (Aneja, Schlesinger and Erisman, 2009). It is worth mentioning that 

interactions occur with other compounds in the atmosphere, to the extent that reductions in SO 2 

and NOx 

are only effective in the reduction of PM 2.5 if carried out simultaneously with NH 3 

reductions (Erisman and 

Schaap, 2004). 

7.4.3 | Aerosols 

Mineral dust, sulphate aerosols, and organic C and black C (BC) aerosols from fossil fuel and biomass burning 

have a significant effect on radiative forcing (Forster et al., 2007). Mineral dust is mainly emitted from deep 

and extensive alluvial flood deposits emplaced during the Pleistocene, for example in the Sahara, East Asia, the 

Arabian deserts, and Central Australia (Prospero et al., 2002). The largest sources are located in the Northern 

Hemisphere, in the so-called ‘global dust belt’ that extends from the west coast of North Africa, through the 

Middle East, into Central Asia. Outside this belt, areas with remarkable persistent dust activity include the 

Great Basin in south-western North America, the Lake Eyre Basinin Australia, some areas of South America 

(predominantly in Argentina), and southern Africa (Prospero et al., 2002) (Figure 6.2). The ‘Red Dawn’ dust 

storm that affected Sydney, Australia in September 2009 is described in Chapter 15. Mineral dust originating 

in the Sahel has been reported to be regularly carried over large areas of the Atlantic and the Caribbean; the 

largest export occurs during years of low rainfall in the source region (Prospero and Lamb, 2003). Although this 

process might have been exacerbated by anthropogenic activities (Prospero and Nees, 1978), recent evidence 

indicates that vegetation cover in the region has not changed substantially in the past 20 years and that, on a 

global scale, dust mobilisation is probably mostly driven by natural events (Prospero et al., 2002). 


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The direct effect of aerosols on the climatic system is mainly through the reflection and absorption of solar 

radiation (Miller and Tegen, 1998). The indirect effect involves the modification of cloud properties (Kaufman, 

Tanre and Boucher, 2002). Greenhouse gases, in contrast, reduce the outgoing thermal radiation to space. 

Differences in lifetime and spatial distribution between greenhouse gases and aerosols are also considerable: 

greenhouse gases have a lifetime of more than 100 years and a homogeneous distribution (Forster et al., 2007), 

whereas aerosols have a lifetime of about a week and a rather heterogeneous distribution (Andreae et al., 

1986). Soil dust aerosols have also been reported to modify the lifetime of some greenhouse gases (Dentener 

et al., 1996). They also provide essential nutrients to ocean ecosystems that may increase the efficiency of the 

ocean’s biological pump and help sequester CO 2 

in the deep ocean (Martin, 1990). This is specially the case of 

iron, which is an important micronutrient for phytoplankton (Falkowski, Barber and Smetacek, 1998). 

Most aerosols are highly reflective, thus raising the albedo of our planet and having a cooling effect. 

However, aerosols containing BC are dark and strongly absorb the incoming sunlight (Kaufman, Tanre and 

Boucher, 2002). This warms the atmosphere and cools the Earth’s surface before a redistribution of the energy 

occurs in the atmosphere column (Ramanathan and Carmichael, 2008). Black C alters the radiative forcing 

through different processes: (i) the presence of BC in the atmosphere above surfaces with high albedo such 

as snow or clouds may cause a significant positive radiative forcing (Ramaswamy et al., 2001); (ii) BC aerosols 

deposited on snow may promote melting (Warren and Wiscombe, 1980; Hansen and Nazarenko, 2004); 

and (iii) BC influences evaporation and cloud formation by modifying the atmosphere’s vertical temperature 

gradient (Ackerman et al., 2000; Raufman and Fraser, 1997). However, the exact radiative forcing depends on 

how BC is mixed with other aerosol constituents (Jacobson, 2001). 

Carbonaceous aerosol emission inventories suggest that approximately 34-38 percent of these emissions 

come from biomass burning sources, the remainder from fossil fuel burning sources (Forster et al., 2007). 

Fossil-fuel-dominated BC emissions are approximately 100 percent more efficient warming agents than 

biomass-burning-dominated plumes (Ramana et al., 2010). The type of smoke is also largely influenced by 

the type of biomass being burned (Takemura et al., 2002). In savannah ecosystems, about 85 percent of the 

biomass (mostly grasses) is consumed by flaming during fire events. In forest fires this value decreases to 50 

percent or less, as the flaming stage is followed by a long, cooler smouldering stage in which the thicker wood, 

not completely consumed, emits smoke composed of organic particles without BC (Takemura et al., 2002). 

Black C is thus mostly emitted during the hot, flaming stage of the fire (Kaufman et al., 2002). The intense 

surface heating caused by fires can further cause a rapid uplift of heated air, known as pyro-convection, which 

can considerably disturb the chemical conditions in the free and upper troposphere and, in some cases, in the 

stratosphere (Monks et al., 2009). Aerosols from fires are more likely to be injected at higher altitudes and are 

likely to experience long-range transport. Aerosol emissions from large boreal fires in Alaska and Russia have 

been shown to be transported very efficiently over long distances (Damoah et al., 2006; Petzold et al., 2007). 

7.5 | Soil change and water quality regulation 

Soils provide a biogeochemically activated filtration and cleaning service that transforms or retains materials 

deposited at the land surface. These materials include not only nitrogen and phosphorous, elements from grey 

water used for irrigation, and acidic compounds, but also inorganic and organic toxins. If the capacity of the 

soil to retain, transform or filter these materials is exceeded, there can be severe environmental consequences 

for water quality. Soils also adversely impact the provision of clean water through erosion into water courses, 

through salinization and through redox cycling and the release of metals such as arsenic. 


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7.5.1 | Nitrogen and phosphorous retention and transformation 

By increasing fertilizer production and crop N fixation, human activities have doubled nitrogen (N) fixation 

from the atmosphere during the last century. Half (210 Tg N yr -1 ) of global nitrogen fixation (413 Tg N yr -1 ) is 

human-driven (Fowler et al., 2013). Mining and erosion have increased the phosphorus (P) flow from land into 

the ocean by at least ten-fold (preindustrial value of 1 to current estimate of 9-32 Tg P yr -1 ; Carpenter and Bennett, 

2011). A recent inventory indicates that approximately 60 percent of the nitrogen fixed by human activities is 

released back into the environment without being incorporated into food or products (Houlton et al., 2013). 

Increases in the release of reactive nitrogen (N) and phosphorus (P) to the environment are associated with 

many significant environmental concerns, including surface water contamination, harmful algal blooms, 

hypoxia, air pollution, nitrogen saturation in forests, drinking water contamination, stratospheric ozone 

depletion and climate change (Bennett, Carpenter and Caraco, 2001; Sutton et al., 2011; Davidson et al., 2012). 

Soils serve as an important regulator of the leakage of this anthropogenic N and P back into the air or to 

surface and ground water, since much of the release occurs from fertilizers or atmospheric deposition. Soil is 

the largest pool of N and P within terrestrial ecosystems (Cole and Rapp, 1981), illustrating the magnitude and 

stability of soil N and P storage. Review of 15N tracer studies reinforces that idea that soils are the strongest 

sink for nitrogen in the short and medium term (Fenn et al., 1998; Templer et al., 2012). Flows of N through 

the landscape and the consequences of excess N can be represented by the N cascade (Galloway et al., 2003). 

Nitrogen and phosphorus removal occurs through plant or microbial uptake, storage in soil organic matter, 

by complexation, and sorption or exchange. Nitrogen is cycled biologically through plant uptake, litterfall and 

microbial cycling, and is stored in organic forms except in areas with substantial rock-derived N (Morford, 

Houlton and Dahlgren, 2011). By contrast, soil P is mainly found in an inorganic form, sorbed or complexed by 

soil minerals and the exchanger. Organic P is a smaller pool in most soils, found in a review of global soil P to 

range from 5-40 percent (Yang and Post, 2011). For N, there are also significant gaseous losses via NOx or NH 3 

and through denitrification as N 2 

or N 2 

O. Storage in soils or perennial plants and conversion into other inert 

forms (N 2 

for N or stable inorganic complexes for P) represent stable sinks that remove N and P from flowpaths 

and the N cascade for a period of time determined by the residence time of those sinks. 

An important service provided by soils is to remove N and P along flowpaths, preventing mobile nitrate 

and phosphate from moving from terrestrial ecosystems into surface waters and groundwater. Global models 

indicate that soils are responsible for the largest portion of landscape N removal - 22 percent of global N removal 

as denitrification - second only to coastal ocean sediments (Seitzinger et al., 2006). Riparian soils or wetlands 

can remove N that has leaked from forests, farms, rangelands or the built environment (Peterjohn and Correll, 

1984), as long as riparian zones are downgradient of the N source (Weller and Baker, 2014). One study indicates 

that replacement of 10 percent of historical riparian buffers could substantially reduce N loading to the Gulf of 

Mexico (Mitsch et al., 2001). 

Phosphorus cycling has important distinctions from N cycling. In particular, the dominant inorganic form 

of phosphorus, orthophosphate, binds strongly to soil particles via sorption or complexation as inorganic P, in 

contrast to nitrate, which is quite mobile. Phosphorus can be displaced under reducing conditions, and thus 

efforts to target N removal may in fact cause unanticipated increases in dissolved P concentrations (Ardón 

et al., 2010). While we do not have a parallel conceptual P cascade, P availability can drive the formation 

of harmful algal blooms, and recent work indicates that joint management of N and P is critical (Conley et 

al., 2009). In efforts to reduce effects on ecosystems and water quality, it is important to consider the soil 

processes involved in removal of both elements and their interactions. 

Perturbations that increase the mobility of N and P may saturate the retention capacity of soils such that the 

ability to remove these elements declines as inputs increase. Disturbances that affect soil structure, rooting 

patterns and organic matter also decrease N and P retention capacity. At the ecosystem scale, N removal 


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capacity declines as N loads increase above a point where N can be taken up by plants and soil processes 

(Aber et al., 1989). While studies illustrate that the rate of N removal does generally decline with increasing N 

inputs (e.g. Perakis, Compton and Hedin, 2005), there are still questions about the ability of soils to retain N 

over time. The saturation point may vary by ecosystem and soil type. For example, wetland ecosystems have 

a tremendous capacity to retain N – a recent meta-analysis indicates that wetland N removal is linear with N 

loading, removing about 47 percent of N inputs even at very high loads (Jordan, Stoffer and Nestlerode, 2011). 

However, recent work on agricultural soils found that N 2 

O production increases with increased N loading 

(Shcherbak, Millar and Robertson, 2014). This reinforces the pattern of decline in capacity of soils to serve as a 

stable N sink under high N inputs, and suggests that efforts to reduce N 2 

O production should target areas of 

high N loads where larger benefits will be seen per unit N. 

The connection between ecosystem services and soil processes is sometimes distant. The benefit of N 

uptake in a riparian soil in Iowa might be most appreciated in distant coastal fisheries. In addition, ecosystem 

services do not turn on or off with the flick of a switch; for example, it may take decades to recover water 

quality after a widespread land use change (Hart, 2003; Howden et al., 2010). Our perspectives about soils and 

ecosystem services should include these distant connections and time lags. 

Removal of N from the cascade has implications for many aspects of human health and well-being (Figure 1; 

Brauman et al., 2007; Compton et al., 2011), and an increasing number of studies are including soil processes in 

ecosystem service assessments and valuation frameworks (De Groot, Wilson and Boumans, 2002; Robinson 

et al., 2013). Soil N and P removal is generally seen as an intermediate service or a supporting or regulating 

service in current ecosystem services classification schemes, as it affects a number of final ecosystem goods 

and services (Boyd and Banzhaf, 2007). 

Impacts of nitrogen on ecosystem services (ES), on the economy and on human well-being have been 

examined in a number of studies (Birch et al., 2010; Compton et al., 2011; van Grinsven et al., 2013). Soil N and 

P storage could have implications for many benefits, including the following: (i) avoidance of consequences 

to ecosystem services provided by freshwater, groundwater and coastal waters from reduced quality for 

swimming, drinking, recreation or fishing; (ii) avoidance of air quality problems associated with N such as those 

affecting human respiratory health or visibility (NOx, NHy); (iii) avoidance of damage from climate change and 

stratospheric ozone depletion (N 2 

O); and (iv) maintenance of soil fertility and ecosystem production (both N 

and P). Eutrophication of coastal areas and associated hypoxia can result in physiological and behavioural 

impacts on important coastal organisms, populations and ecosystems that result in lowered fitness and 

productivity. However, there is a good deal of uncertainty about the economic damages associated with 

coastal eutrophication in many areas (Rabotyagov et al., 2014). Efforts to inform policy should bring together 

ecologists and economists to study the impacts of N and P on ecosystem services all along the cascade. 

7.5.2 | Acidification buffering 

Soil acidity is controlled by both biota (plant roots and microorganisms) and particles (soil minerals and 

organic matter). Production of carbon dioxide, organic matter decomposition, and the excretion of acidic 

compounds by biota increase soil acidity, while binding of acidic compounds to root and particle surfaces, 

as well as mineral weathering, decrease it (Sposito, 2008). Over periods ranging from centuries to millennia, 

while most of the less resistant minerals become depleted through weathering reactions with rainwater and 

subsequent leaching, highly acidic soils are produced naturally. They now occupy about one-third of the icefree 

land area on Earth (Guo et al., 2010), mainly in the humid tropics and in the forested regions of temperate 

zones. 


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Industrial effluents (for example, sulphur and nitrogen oxide gases dissolved in atmospheric precipitation 

or transformed to particles, or acidic wastewaters) and nitrogenous fertilizers, such as urea, are typical 

anthropogenic inputs of acidity to soils. If these two acidic inputs exceed about 15 percent of the capacity of 

soil to neutralize them, acidification increases markedly, with a variety of serious problems arising for both 

plant and microbial growth. The potential for generating polluted runoff or drainage water also increases 

markedly. Over a 20 year period Guo et al. (2010) documented such increases of acidity in Chinese topsoils, 

caused by nitrogen fertilization and acidic deposition. The topsoils investigated showed an average pH 

decrease of 0.50, which is quite serious. Other long-term studies document decadal changes in soil acidity 

that are even larger (Richter and Markewitz, 2001). Acidic deposition is an important problem in China, but 

the acidification caused by nitrogen fertilization was found to be 10 to 100 times greater than that caused by 

acid rain. In the principal double-cropping cereal systems of China (wheat-maize, rice-wheat, and rice-rice), 

nitrogen fertilizer use efficiencies are only 30 to 50 percent. The progressive acidification of topsoil – as well 

as nitrogen pollution of agricultural runoff and drainage – will remain unchecked as long as this low nitrogen 

use efficiency is not addressed. Guo et al. (2010) noted that optimal nutrient-management strategies can 

significantly reduce nitrogen fertilization rates without decreasing crop yield, thus providing benefits to both 

agriculture and water quality. 

7.5.3 | Filtering of reused grey water 

Nearly 80 percent of urban ‘blue water’ becomes wastewater. At about 100 m3yr -1 per household in the 

developed world, wastewater thus represents a rapidly expanding environmental and health challenge, 

particularly in urban centres. The ecological footprint of untreated wastewater is unsustainable even in 

regions where water is plentiful (e.g. South East Asia), as it may either increase nutrient loads in rivers and 

coastal regions or represent a direct hazard to human health. By contrast, arid regions increasingly rely on 

treated wastewater for irrigation, often practiced with little consideration of long-term impacts on the soil, 

hydrology and ecology of the producing area. The sustainability of this coupled agro-urban hydrological cycle 

hinges on proper management to mitigate adverse impacts of long-term wastewater use and avoid potential 

collapse of soil ecological functions. Various studies (e.g. Bond, 1998; Assouline and Narkis, 2013) have shown 

that, over the long term, even irrigation with wastewater results in significantly increased soil ESP that can 

adversely impact soil structure and hydraulic properties. In the absence of proper regulation, irrigation with 

wastewater may pose a range of human health and other ecological risks associated with introduction 

of pathogenic microorganism into the soil and crop (del Mar et al., 2012). The sustainable management of 

wastewater irrigation requires new management strategies including water source mixing, proper selection 

and rotation of crops, and avoidance of sensitive soils. 

7.5.4 | Processes impacting service provision 

Trace elements 

Elevated concentrations of potentially toxic trace elements can affect provision of the services that depend 

on soils. Trace elements – such as arsenic, cadmium, chromium, lead, mercury, and selenium – naturally occur 

in low quantities within soils. They may also be introduced and concentrated through anthropogenic activities 

like waste disposal, fertilizer and pesticide application, and atmospheric particulate emission and deposition 

(Sparks, 2003; Pierzynski, Vance and Sims, 2005). Even when at low concentrations in soils, they can have 

pronounced impacts on water quality. This is particularly the case where the capacity of soils to store trace 

elements is exceeded or where there are changes in the soil chemical, physical and/or biological environment 

that influence the partitioning of trace elements between the solid and aqueous phases. 


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The concept of the critical load of a specific trace element enables a precautionary assessment of the risks 

its input causes to food quality and of the eco-toxicological effects on organisms in soils and surface waters 

(Lofts et al., 2007; de Vries et al., 2013b). The critical load of trace elements is defined as “the load resulting 

at steady state in a concentration in a compartment (e.g. soil solution, plant, fish) that equals the critical 

limit for that compartment” (Lofts et al., 2007; de Vries et al., 2013b). The critical limit is a receptor-specific 

concentration below which significant effect on the receptor is assumed not to occur (Lofts et al., 2007). The 

concept of critical loads – specifically the critical loads of acidity − was key in gaining acceptance of the need for 

reduction of atmospheric deposition of N and S (Section 4.4.1 above). However, the usefulness of the concept 

of critical loads of trace elements in international negotiations aimed at reducing trace element deposition 

is not equally evident. This is mainly owing to two factors that distinguish trace elements from the case of 

acidity and acid rain: (i) the time needed for a specific trace element in a specific scenario to attain steady 

state is much longer than for N and S; and (ii) other changes in the environment, notably acidification, may 

have a greater influence on the exposure and effects of a specific trace element than the particular amount 

entering the system (de Vries et al., 2013b). In fact, problems associated with trace elements in soils are 

commonly exacerbated by changes in land use that alter environmental conditions and increase the potential 

for exposure to trace elements through food and water consumption. Because of this, in addition to applying 

the concept of critical loads, the assessment of the future risks of trace elements needs to employ dynamic 

models (de Vries et al., 2013b). 

Salinity 

Salinization of soil and water resources remains a chronic problem in many parts of the world, mostly in 

arid regions where evapotranspiration exceeds rainfall. The increased frequency of extreme climate events 

(droughts, intense rainfall events) together with the expansion of irrigated agriculture are expected to increase 

the range of soils affected by salinity. 

In addition to the effects of hotter and drier climate patterns, the primary causes of salinity risk include: 

(i) increasing salt loads due to use of marginal water sources such as waste water; (ii) over exploitation of 

coastal aquifers and related sea water intrusion (Várallyay, 1994); (iii) overpumping and degradation of slowly 

replenishing inland aquifers (Ogallala); (iv) sea level rise impacting coastal wetlands (e.g. Mexico pacific 

coastline); (v) mismanagement of rapidly expanding irrigation in arid regions, particularly inadequate leaching 

and drainage and (vi) clearing of perennial vegetation in landscapes with significant salt stores in soils and 

deeper regolith. 

One solution is to reduce the salt content of irrigation water through desalination. Recent advances in 

desalination techniques have resulted in a dramatic reduction in costs. Irrigation experiments with desalinated 

water show substantial increase in yield with less water used and less salt leaching to groundwater resources. 

However, the use of desalinated water requires careful management to avoid soil and ecological damage (e.g. 

clay dispersion) due to irrigation with extremely pure water (Yermiyahu et al., 2007; Tal, 2006). 

Erosion 

Intensification of agriculture, changes in rainfall patterns with more intense rain events, and potentially 

more compacted soil surfaces may all contribute to increased rates of surface soil erosion. In addition to 

the removal of the top layer of productive soil and the incision of stream channels, the potential increase in 

soil transport to surface water may cause a cascade of adverse effects downstream. Pimentel et al. (1995) 

list impacts on stream and lake ecology, dam siltation and effects on waterways, and of course, potential 

for enhanced pollution by agrochemicals and colloid-facilitated transport of phosphorous and carbon. Soil 

erosion is also linked to climate change as it mobilizes large amounts of soil organic carbon (SOC). Since the 

industrial revolution and associated land use changes, SOC has been estimated to contribute 78±12 Gt of C 

to the atmosphere, of which about one-third is due to accelerated erosion and two-thirds to mineralization 

(WMO, 2005). 


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The WMO (2005) report estimates that 25 percent of African soils are prone to risk of water erosion (excluding 

deserts that comprise about 46 percent of the African land surface), and that 50 percent of cropland in Australia 

is susceptible to water erosion. Drier conditions associated with future climate extremes (droughts) may 

limit rates of soil carbon accumulation and reduce soil aggregation, thereby enhancing vulnerability to wind 

erosion. WMO (2005) estimate that about 22 percent of the African land surface is prone to wind erosion, and 

15 percent of the cropland in Australia. A host of soil conservation strategies for combating land degradation 

due to soil erosion also offer co-benefits such as enhanced water storage in the soil profile (Pimentel et al., 1995; 

Troeh, Hobbs and Donahue, 1991). Eroded landscapes may take centuries to millennia before their abilities to 

provide quality ecosystem services are restored. 

7.6 | Soil change and water quantity regulation 

Soil moisture regulation of precipitation 

Soil moisture acts as a buffer for precipitation anomalies. As long as the soil is not saturated, it can reduce 

the direct impact of flooding. Similarly, soil moisture acts as a buffer against dry anomalies in the onset of 

meteorological droughts, before soil moisture or streamflow droughts are noticeable. However, if preevent 

soil moisture is anomalously wet or dry, these same properties can also lead to significant flooding 

and droughts even where precipitation is not abnormally high or low. For these reasons, the monitoring of 

soil moisture conditions (as well as of snow and groundwater) is valuable for the forecasting of floods and 

droughts (e.g. Koster et al., 2010b; Fundel, Jörg-Hess and Zappa, 2013; Orth and Seneviratne, 2013; Reager, 

Thomas and Famiglietti, 2014). 

In addition to effects related to the buffering or persistence of soil moisture, several studies suggest that 

soil moisture also affects the regional water cycle through impacts of evapotranspiration on precipitation (e.g. 

Beljaars et al., 1996; Koster et al., 2004; Seneviratne et al., 2010; Taylor et al., 2012). However, the underlying 

feedbacks, including their sign, are strongly model-dependent (e.g. Koster et al., 2004; Hohenegger et al., 

2008). Also observational studies diverge with respect to inferred soil moisture-precipitation feedbacks. 

Some suggest the presence of positive (temporal) feedbacks while others identify mostly negative (spatial) 

feedbacks (Findell et al., 2011; Taylor et al., 2012). In addition, causality is very difficult to establish based on 

observations (e.g. Salvucci, Saleem and Kaufmann, 2002). Precipitation persistence could, for example, lead 

to some confounding effects (Guillod et al., 2014). Overall, effects of soil moisture on precipitation are still 

uncertain. 

Human land and water use strongly affects soil moisture variations and the resulting land water balance, 

for instance through irrigation (Wisser et al., 2010; Wei et al., 2013) or other changes in agricultural practices 

(Davin et al., 2014; Jeong et al., 2014). These effects are generally not considered in present day climate models, 

although they could substantially affect soil moisture and hydrological drought projections, including 

feedbacks to the atmosphere. 

7.6.2 | Precipitation interception by soils 

Together with vegetation, soils help to regulate water quantity by intercepting water, reducing floods 

and maintaining the soil moisture buffer. Precipitation arriving at the Earth’s surface can be intercepted by 

vegetation canopies and returned directly to the atmosphere through evaporation, never reaching the soil 

moisture pool. Typically, trees can intercept 25-50 percent of precipitation and shrubs 10-25 percent, while 

interception by grass is significantly less (Calder, 1999). The rest of the precipitation arrives at the soil surface, 

the characteristics of which control the partitioning between what infiltrates and what runs off into surface 

water. In a recent meta-analysis, Jarvis et al. (2013) have shown that K is largely dependent on bulk density, 

organic carbon content and land use. This has important consequences for ecosystem service delivery by soils, 


194

as it indicates that management and land-use change will affect the soil infiltration service temporally as well 

as spatially. This analysis by Jarvis et al. (2013) corresponds to an increasing number of studies that show the 

importance of vegetation in determining soil K values on similar soils. 

Because of their large root systems, trees in particular create conduits for conducting water into soil. Both 

dead and living roots can create flow networks. Beven and Germann (1982) cited work suggesting that as much 

as 35 percent of the volume of a forest soil may contain macropores formed by roots. Chandler and Chappell 

(2008) demonstrated that K was highest near the trunk of single oak trees and decreased toward the edge of 

the canopy. The ratio of K geometric mean values under the tree at 3 metres from the trunk to the adjacent 

pasture was 3.4 times higher, similar to results compiled from the literature in the same paper. Gonzalez-Sosa 

et al. (2010) presented conductivity data for a range of land use types in France, with trees being generally 

higher, and crops and pasture lower for the same soil. In the tropics, deforestation results in a major reduction 

in infiltration, whether the forest is recently cleared or has been turned into pasture (Zimmermann, Elsenbeer 

and De Moraes, 2006). 

Soil macrofauna - worms, ants and termites etc. - also play an important role in determining infiltration at 

local scales (Beven and Germann, 1982; Lal, 1988), and perhaps also regionally and globally given the prevalence 

of these organisms. There are typically two modes of macrofauna action impacting hydraulic properties. The 

first is the creation of burrows forming macropores; the other is the turnover of soil and aggregation which 

impacts infiltration and water retention, generally increasing both. Soils might offer potential for slowing 

water movement across landscapes under certain precipitation conditions (Marshall et al., 2009). However, 

once runoff is generated and large quantities of precipitation fall, the role of soils is likely to be less important. 

Above a certain threshold, massive floods can occur in almost any landscape 

Although often cited as an important ecosystem service, the impact of land use on altering flood risk 

remains hard to quantify with any precision (Pattison and Lane, 2011). The link between land management 

and flood risk is complex and scale dependent as conceptualized by Bloschl et al. (2007). Many studies have 

demonstrated how land or soil management impact infiltration and runoff generation at the plot to hillslope 

scale (Wheater and Evans, 2009). These tend to be local effects in temperate zones, but can be large scale in the 

tropics. Beven et al. (2008) found a distinct land use signal hard to detect, and also pointed out that “adequate 

information about past land management changes and soil conditions is not readily available but will need 

to be collected and made available in future for different land use categories if improved understanding of the 

links between runoff and land management is to be gained and used at catchment scales.” 

7.6.3 | Surface water regulation 

Soils provide a maintenance service that contributes to the regulation of base flow and water supply in 

rivers. Groundwater, lakes and soil drainage all play a role in setting base flow in surface waters (Price, 2011). 

Groundwater dominates in the lowlands, but soil drainage dominates upland catchments. Changes to the 

hydraulic characteristics of upland catchments and to the quantity of water stored by soils will have distinct 

implications for water supply downstream. Again, the soil water retention characteristics and hydraulic 

conductivity play a crucial role in the regulation of drainage. 

7.7 | Soil change and natural hazard regulation 

Soil and its characteristics (depth, hydro-mechanical properties, mineralogy, ecological function, and 

position in the landscape) play an important role in several natural hazards including: landslides, debris flows, 

floods, dam failure, droughts, shrink and swell damage to roads and infrastructure, and more. The United 

Nations International Strategy for Disaster Reduction (UNISDR, 2009) defines a natural hazard as a “natural 


195

process or phenomenon that may cause loss of life, injury or other health impacts, property damage, loss 

of livelihoods and services, social and economic disruption, or environmental damage”. Projected human 

population expansion, agricultural intensification, and greater human presence and infrastructure in 

mountainous regions combined with projected changes in climate extremes (IPCC, 2012) are expected to 

jointly contribute to enhanced vulnerability to soil-mediated natural hazards (Figure 7.9). The extent of the 

vulnerability and exposure to a particular type of hazard vary considerably among regions (ESPON, 2013). 

For example, floods may increase in flat terrains with increasing mean precipitation or rapid snowmelt, and 

landslides may become more common in mountainous areas with changes in the seasonality and intensity of 

rainfall (Huggel, Clague and Korup, 2012). 

Figure 7.9 A conceptual sketch of how vulnerability, exposure and external events (climate, weather, geophysical) contribute to the 

risk of a natural hazard. Source: IPCC, 2012. 

The past few decades have been marked by an increase in the frequency and magnitude of damages caused 

by soil-climate related hazards such as landslides (Figure 7.10, FAO 2011). In part this increase may be simply 

attributed to more timely and accurate reporting, and also to deeper human penetration into soil-hazard 

prone regions, facilitated by increases in mobility and personal wealth (Keiler, 2013; Papathoma-Köhle et al., 

2015). The reports of EM-DAT (http://www.emdat.be/publications) provide a global perspective of all aspects of 

natural disasters and their human and economic impacts. The 2013 EM-DAT 1 report estimates global damages 

by natural hazard attributed to hydrological and geophysical causes (most closely related to soil) in excess of 

US$ 60 billion, with impacts on the lives of 40 million people in 2013 alone. It is instructive to place the various 

natural hazards in their soil-human-climate context to enable general inferences and detection of future 

trends with global change (population growth, land use, and climate change). 

1 http://www.emdat.be/publications 


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10-year frequency 

Ten-year frequency of landslides and associated mortality in Asia 

Mortalities 

120 

7000 

100 

80 

60 

40 

Frequency 

Mortalities 

6000 

5000 

4000 

3000 

2000 

20 

1000 

0 

1950 - 1959 

1960 - 1969 

1970 - 1979 

1980 - 1989 

1990 - 1999 

2000 - 2009 

0 

Figure 7.10 Trends in landslide frequency and mortality on Asia. Source: FAO, 2011; EM-DAT, 2010. 

7.7.1 | Soil landslide hazard 

The depth of the soil mantle forming over mountainous topography reflects a natural balance between 

soil production and soil erosion processes (Trustrum and De Rose, 1988; Heimsath et al., 1997). The primary 

soil removal process in mountainous regions is landsliding, driven by the topographic relief and triggered by 

climatic forcing such as rainfall or snowmelt (Iverson, 2000; Larsen, Montgomery and Korup, 2010; Kawagoe, 

Kazama and Sarukkalige, 2009) or by earthquakes (Huang and Fan, 2013). Landslide damage is costly: Sidle 

and Ochiai (2006) estimated the direct costs associated with rebuilding or replacing infrastructure at several 

billion dollars per year, even without considering indirect costs related to construction and temporary loss of 

site functionality. Similar estimates have been made just for Europe (Papathoma-Köhle et al., 2015). 

Rainfall is the most common trigger for shallow landslides (Iverson, 2000). The strong relationship between 

rainfall intensity-duration and landslide triggering conditions has prompted the use of rainfall characteristics 

for early warning (Guzzetti et al., 2008; von Ruette, Lehmann and Or, 2014). The observed increase in 

precipitation variability and in extreme events attributed to climate change has been linked to the observed 

increase in landslide frequency in mountainous regions (Huggel, Clague and Korup, 2012). The recent IPCC 

report (IPCC, 2012) lists evidence for the contiguous United States confirming statistically significant increases 

in heavy (upper 5 percent) and very heavy (upper 1 percent) precipitation of 14 and 20 percent, respectively. 

Moreover, evidence from Europe and the United States suggests that the relative increase in precipitation 

extremes is larger than the increase in mean precipitation. 

Schmidt and Dikau (2004) found that climatic scenarios representing unstable conditions of transition 

from more humid to a dryer climate produced the highest slope instabilities. Soil hydraulic properties play an 

important in imparting mechanical sensitivity. Indeed, the soil plays multiple roles in the landslide hazard, not 

only as the mass that slides down the slope, but also through its own mechanical strength and through its 

modulation of local hydrology via infiltration capacity, base flow, macropore flow and ground cover (Iverson, 

2000; Sidle and Ochiai, 2006; Lehmann and Or, 2012). The partitioning of precipitation between infiltration, 

overland flows and base flows is critical to the loading of the soil and to the ultimate soil failure. The mechanical 

reinforcement by plant roots helps to stabilize the soil mantle (Abe and Ziemer, 1991; Schwarz, Cohen and Or, 

2012), and bulk soil mechanical and hydraulic properties affect the susceptibility to failure. 


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Recent widespread drought-induced forest die-offs highlight how climate change could accelerate forest 

mortality. This has potential consequences for the carbon cycle and for ecosystem services (Anderegg et 

al., 2013). Through loss of root reinforcement, die-off may also increase landslide hazard. Rapid landslide 

processes have also been observed in Southeast Asia and the Western Pacific where large tropical cyclones 

induce numerous landslides and remove significant amounts of soil and particular carbon through the river 

systems to the ocean (Hilton et al., 2008). These extreme tropical precipitation events are likely to increase in 

frequency and magnitude (Huang et al., 2013). 

7.7.2 | Soil hazard due to earthquakes 

Keefer (2002) provides a historical overview of the study of earthquake-induced landslides. These are often 

extensive in their size and occurrence and cause more significant damage than hydrologically-induced shallow 

landslides. For example, the 2008 Wenchuan earthquake in Sichuan province in China triggered more than 

60 000 landslides over an area of 35 000 km 2 causing about one-third of the total number of fatalities in the 

earthquake disaster (Huang and Fan, 2013). In addition to the direct damages, the Wenchuan earthquake 

induced an unprecedented number of secondary geohazards such as heightened subsequent landslide 

frequency, causing river damming and consequent floods as well as debris flows. The links between seismic 

activity and landslide characteristics were systematically investigated by Malamud et al. (2004) based on 

landslide inventory data of landslide size-frequency distribution in the affected landscape. These analyses are 

useful for deriving large-scale soil erosion rates enhanced by seismic activity. Erosion rates in active subduction 

zones are around 0.2–7 mm yr -1 . Hazard schemes often classify earthquake-induced landslides as ‘geophysical’ 

or ‘dry’ events to indicate they do not require water for mass movement initiation, unlike hydrological ‘wet’ 

landslides. 

On March 11, 2011, a seaquake followed by an enormous tsunami and by the destruction of the Fukushima 

Atomic Power Plant, Japan, brought about additional soil changes such as liquefaction, tsunami sedimentation 

and radio isotope contamination, all of which affected the local population. Liquefaction brought about by the 

earthquake occurred mainly on soil-banked lands or soil-dressed lands, causing extreme damage to housing 

and structural facilities. The tsunami carried massive deposits from the bottom of the sea onto farmlands 

along the seashore. This sedimentation contained considerable quantities of arsenic (Kozak and Niedzielski, 

2013). The explosion of the atomic power plant resulted in soil contamination (mainly with Cesium 137) of an 

area as large as 800 square kilometres (Steinhauser, 2014; Itoha et al., 2014). Cleaning these contaminants is 

vital before the population can return. More broadly, although a variety of soil hazard regulation techniques 

have been developed (Gasso et al., 2013; Delgado et al., 2011; Esteves et al., 2012) there is a need for both more 

research and more regulation related to soil hazards than hitherto. 

7.7.3 | Soil and drought hazard 

Droughts limit primary production and thus the accumulation rates of organic matter. Reduced 

accumulation rates contribute to soil vulnerability to water and wind erosion. Recent meso-scale strategies for 

combating drought damage and reducing risk in agro ecosystems have proposed landscape-scale vegetation 

management. This can, for example, take the form of patches or bands of perennial vegetation to promote 

feedbacks that are conducive to recycling of water vapour, soil moisture and nutrients (Ryan, McAlpine and 

Ludwig, 2010). An often ignored consequence of prolonged drought and soil water depletion is soil subsidence 

and related damage to buildings and infrastructure (Corti et al., 2011). Corti et al. (2011) presented a systematic 

study of damage costs from drought-induced soil subsidence applicable across different climate regimes. The 

primary variables include drought severity, soil type (shrink/swell properties), land use, and vegetation. 

Prolonged droughts and drier climate patterns accentuate damages due to soil shrink/swell properties. 

The insurance industry reports that damage to infrastructure often peaks following extreme drought 

events, especially in the densely built up regions of Europe and United States. Dry climate also induces other 


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phenomena such as the onset of massive dust storms. Dust storms can arise either from destabilization of 

vulnerable surface soils (the Dust Bowl), or from the drying of lake beds, or from desertification and loss of 

vegetation and similar soil destabilizing activities over large scales. The rates of wind erosion associated with 

sand storms may exceed 100 mm topsoil yr -1 in sensitive regions in the Sahel. Prolonged exposure is known to 

pose respiratory health hazards to human population. 

7.7.4 | Soil and flood hazard 

Agricultural intensification has been linked to alteration of runoff mechanisms and to increased risk and 

burden of floods (Marshall et al., 2014). Some of the primary changes in land management documented in 

the United Kingdom and elsewhere that affect soils include: heavy traffic contributing to soil compaction, 

tillage operation and consequent loss of soil structure, the formation of larger fields, choice of cover crops in 

rainy seasons, and increased livestock densities (O’Connell et al., 2007). However, establishing rigorous causal 

links between changes in land management practices, local runoff generation and catchment scale flood 

behaviour remains a challenge (Ewen et al., 2013). Nevertheless, mounting evidence suggests that soil and 

land management contribution to flood risk is not limited to management of lowland agricultural regions. 

Management of upland soils and related impacts on runoff generation mechanisms cascade and also have 

impacts on flood risk downstream (Wheater and Evans, 2009; Marshall et al., 2009). A recent review by Hall et 

al. (2014) on flood trends in Europe (including climatic effects) confirms the important role of land use changes 

(urbanization, afforestation, etc.) as key factors in modifying large scale flood risk. Some of the strategies for 

reducing flood risk include afforestation in upland catchments (Ewen et al., 2013), creation of retention basins, 

and adding floodplains by lowering levees (Hall et al., 2014). 

7.7.5 | Hazards induced by thawing of permafrost soil 

Permafrost is perennially frozen soil remaining at or below 0°C for at least two consecutive years (Brown et 

al., 1998). Permafrost regions occupy about 24 percent of the exposed land area in the Northern Hemisphere 

and in some high mountainous regions (UNEP, 2012). Expected thawing of permafrost is projected to induce 

alterations in soil hydrology and biological activity, and to have an impact on the global carbon cycle (Schuur 

et al., 2008). In addition, the thawing of permafrost is expected to change vegetation species and reshape 

many ecosystem functions. The mechanical weakening of the previously frozen soil is likely to result in 

foundation settling, with damage to buildings, roads, pipelines, railways and power lines (Nelson, Anisimov 

and Shiklomonov, 2001; Jorgenson, Shur and Pullman, 2006). Estimates of infrastructure repair in Alaska up 

to 2030 are in the range of US$ 6 billion (UNEP, 2012). Changes in mean temperature and snow cover also 

affect sensitive permafrost in high mountains, and contribute to a higher risk of landslides and avalanches 

(Gruber and Haeberli, 2007; Harris et al., 2009). Schoeneich et al. (2011) present an extensive report and case 

studies, largely from the European Alps, on various slope movement hazards (landslides, rock fall, and debris 

flow initiation) associated with degrading permafrost. Evidence suggests accelerated erosion rates of the 

thawed permafrost, especially along coastlines and rivers banks as documented by Schreiner, Bianchi and 

Rosenheim (2014) and Vonk et al. (2012), with subsequent transport of the carbon-rich sediment through river 

systems to the ocean. 

7.8 | Soil biota regulation 

Soil biodiversity is vulnerable to many anthropogenic disturbances, including land use and climate change, 

nitrogen enrichment, soil pollution, invasive species and the sealing of soil. A recent sensitivity analysis 

revealed that increasing land use intensity and associated soil organic matter loss are placing the greatest 

pressure on soil biodiversity (Gardi, Jeffery and Saltelli, 2013). Numerous studies report soil biodiversity declines 

as result of the conversion of natural lands to agriculture (Bloemers et al., 1997; Eggleton et al., 2002; Dlamini 


199

and Haynes, 2004), and as a result of agricultural intensification (Mulder et al., 2005; Postma-Blaauw et al., 

2010; De Vries et al., 2013a). In particular, studies show larger bodied soil animals, such as earthworms and 

termites are especially vulnerable, but intensive land use can also reduce the abundance and variety of species 

of nematodes, mites and collembolans. 

Climate change also poses a considerable threat to soil biodiversity through direct effects of warming and 

altered precipitation (e.g. drought and flooding) on the availability of moisture in soil (Bardgett et al., 2008). 

Indirect climate change effects of warming and elevated atmospheric carbon dioxide may also have an impact 

on the quantity and quality of organic matter in soil (Blankinship, Niklaus and Hungate, 2011; van Groenigen et 

al., 2014). Although poorly understood, predicted increases in the frequency of erosive rainfall events (Nearing 

et al., 2005) and climate-induced shifts in land use (Mullan, 2013) could pose a considerable future threat to soil 

biodiversity. Other threats to soil biodiversity include nitrogen enrichment, which negatively impacts soil fungi 

(Treseder, 2008), soil sealing, which effectively stops the natural functioning of soil (Gardi, Jeffery and Saltelli, 

2013), and invasive species, which affect native soil biodiversity through a range of mechanisms, including 

altered resource supply, competitive interactions and predation, and physical and chemical modification of 

the soil environment (Wardle et al., 2011). 

Although it is well known that soil organisms play key roles in many ecosystem processes, our understanding 

of the functional consequences of belowground diversity loss is limited, at least compared to what is known 

about aboveground losses (Cardinale et al., 2012). Recent synthesis of experimental studies on soil diversityfunction 

relationships indicate that diversity effects on processes of nutrient and carbon cycling are highly 

variable, but effects of species loss are most pronounced at the low end of the diversity spectrum (Nielsen 

et al., 2011). There is also a general consensus that changes in the functional composition of belowground 

communities, rather than species diversity per se, are of most importance for ecosystem functioning (Nielsen 

et al., 2011). Consistent with this, laboratory studies with low numbers of species have shown the functional 

composition of soil macrofauna communities to be a better predictor of litter decomposition than species 

richness (Heemsbergen et al., 2004). The selective removal of different groups of soil organisms has been shown 

to impair soil functioning (Wagg et al., 2014). Likewise, a recent cross-biome field experiment showed that the 

loss of key components of the decomposer communities consistently slowed rates of litter decomposition and 

carbon and nitrogen cycling, indicating negative effects of diversity loss on soil functions (Handa et al., 2014). 

A field-based study of different sites across Europe also showed that changes in soil food web composition 

resulting from intensive agriculture consistently strongly affected processes of carbon and nitrogen cycling 

(De Vries et al., 2013a). At one site, high intensity management reduced the resistance and resilience of the soil 

food web to drought, increasing soil carbon and nitrogen loss as greenhouse gases and in leachates (De Vries 

et al., 2012a, 2012b, 2012c). 

Changes in soil biodiversity can also modify vegetation dynamics, both directly through associations of 

symbionts and pathogens with plant roots, and indirectly, by modifying nutrient availability to plants (van 

der Putten et al., 2013). For example, mycorrhizal fungi, which form symbiotic associations with roots of most 

plant species and are very vulnerable to soil disturbances, can enhance plant species diversity by relaxing plant 

competition intensity and promoting more equitable distribution of resources within the plant community 

(van der Heijden, Bardgett and van Straalen, 2008). Also, plant diversity and productivity have been shown, 

in some situations, to be positively related to arbuscular mycorrhizal fungal diversity due to more efficient 

use of soil phosphorus (van der Heijden et al., 1998). Soil pathogens, which cause considerable problems for 

agricultural crops, have also been shown to impact vegetation dynamics in natural settings, by suppressing 

the growth of their host plant species more than their neighbours, thereby contributing to vegetation change 

(Bever, Westover and Antonovics, 1997; Packer and Clay, 2000; Klironomos, 2002). The spread of invasive 

plant species has also been linked to release from their natural soil enemies in their new territories, giving 

the invasive plant a competitive edge over native species. This often leads to declines in plant diversity and to 

shifts in the functioning of the soil (Wardle et al., 2011; van der Putten et al., 2013). 


200

Although poorly explored, diversity changes in soil are likely to impact soil physical properties, with 

consequences for ecosystem services related to soil formation and water regulation. Diversity effects on soil 

physical properties have not been explicitly studied, but they are likely to be important given the potential for 

different groups of soil organisms to differentially impact soil structure through different routes. For example, 

fungi promote soil aggregate stability through the physical enmeshment of soil particles by their extensive 

networks of mycelia, whereas bacteria produce metabolic products, mainly polysaccharides, which bind 

soil particles together (Hallett et al., 2009). Mycorrhizal infection can also influence soil aggregate stability 

through physical enmeshment of soil particles by their extensive networks of mycelium, but also through the 

binding of soil particles via the production of extracellular polysaccharides and proteins, including the protein 

glomalin, which alters the wetting behaviour of soil (Rillig and Mulley, 2006). Finally, soil animals, especially 

ecosystem engineers such as earthworms and termites, impact soil structure by creating macropores and 

channels, thereby improving water movement through soil (Bardgett, 2005). 

While evidence is mounting that shifts in soil biodiversity resulting from human activities have significant 

consequences for ecosystem functions and the services that they underpin, there is still much to be learned. 

The mechanisms by which soil biodiversity change can impact ecosystem are enormous, involving a range of 

ecological and evolutionary processes at different spatial and temporal scales, and links between aboveground 

and soil communities. Moreover, impacts of soil biodiversity change on soil functions are likely to be context 

dependent, varying with soil abiotic properties and vegetation type. Unravelling this complexity in order to 

make better predictions about the consequences of soil biodiversity change for the services that ecosystems 

provide is a major challenge. 

7.9 | Soils and human health regulation 

The linkage between soils and human health is increasingly being recognized (Abrahams, 2006, 2013; 

Baumgardner, 2012; Brevik and Burgess, 2013; Jeffrey and van der Putten, 2011; Oliver, 1997). A central 

understanding is that soils form an integral link in a holistic view of human health that includes physical, 

mental and social dimensions. The soil acts as a natural filter, it can kill off pathogens, it can biodegrade 

organics and, in general, it does a wonderful job of protecting us from human health threats. However, soil is 

not able to protect itself against all the insults it is subject to on a regular basis. 

Soils aid in the regulation of human health. They do this by keeping in check, or balancing, the beneficial 

versus deleterious concentrations of elements and moderating disease-causing organisms. For example, soils 

regulate human health by impacting the nutrient quality or nutrient density of foods. Too little of an essential 

nutrient in soil can lead to human diseases such as Keshan disease caused by selenium deficiencies in the 

human diet (Chen, 2012). Conversely, health problems can be caused by an excess amount of organics or trace 

elements such as the arsenic released by soils into the drinking and irrigation waters of Bangladesh (Khan, 

Hamra and Mu, 2009) (Section 7.3). 

Soil is a natural source of radiation that can adversely affect human health, and soil can also affect human 

health by directly interacting with people. One example is the disease of podoconiosis or Mossy Foot disease 

(Mossy Foot Project, 2014). Mossy Foot disease affects about 5 percent of the population in highland tropical 

areas with volcanic soils and lots of rainfall. These soils are rich in silicates that can penetrate the skin of 

susceptible people as they go barefoot about their daily business. Soils can also act as a reservoir of all kinds 

of introduced materials that can impact human health. The dioxin at Love Canal in New York, United States 

is a classic example (Silkworth, Culter and Sack, 1989). There are large quantities of industrial and agricultural 

products and by-products added to soil every year that have the potential to impact human health. 


201

Vast numbers of people, primarily women, infants and children, are afflicted with trace element deficiencies 

(notably Fe, I, Se, and Zn), mostly in the resource-poor countries of the developing world. A diet with low 

boron (B) has been found to lead to a number of general health problems and to increase cancer risk. The 

most common symptoms of B deficiency include arthritis, memory loss, osteoporosis, degenerative and 

soft cartilage diseases, hormonal disequilibria and a drop in libido (Scorei and Popa, 2010). According to one 

hypothesis, the low cervical cancer incidence in Turkey correlates with its B-enriched soil (Simsek et al., 2003). 

Indeed, the ingestion of B via drinking water prevents cervical cancer risk (Korkmaz et al., 2007). In a survey 

in northern France, exposure to high levels of boron (>0.3 mg L -1 ) in the drinking water was associated with 

a significantly lower mortality rate as compared to that of a low-boron reference area (Yazbeck et al., 2005). 

Silicon is the second most abundant element in the Earth’s crust. Dietary silicon intake is positively 

associated with bone mineral density in men and premenopausal women of the Framingham Offspring cohort 

(Jugdaohsingh et al., 2004). Silicon is bound to glycosaminoglycans and has an important role in the formation 

of cross-links between collagen and proteoglycans (Carlisle, 1976). In vitro studies have demonstrated that 

silicon stimulates type 1 collagen synthesis and osteoblast differentiation (Reffitt et al., 2003). 

Many physicians have believed that zinc deficiency is a rare occurrence in Japan. Nevertheless, One study 

found many zinc-deficient patients at a clinic in Japan since 2002 (Kurasawa, Kubori and Okuizumi, 2010). 

Their complaints were anorexia, general fatigue, impaired sense of taste, burning mouth, various types of skin 

lesion, delayed wound healing and emotional instability. 

Based on dietary intake recommendations, subclinical or marginal Mg deficiency (50 percent to

affect the health of humans and animals and entire ecosystems. The same techniques that are available to 

map the human microbiome can also be applied to map the soil microbiome. We are thus on the verge of 

understanding what constitutes a healthy soil microbiome and how a degraded or unhealthy soil microbiome 

may affect our food production and overall human health. 

7.10 | Soil and cultural services 

The soil is one of the main sources of information on the prehistoric culture of humankind. Indeed, soil 

is an excellent medium for preserving artefacts. Different soil types have particular characteristics to 

preserve remains. For instance, in permanently or seasonally wet soils, the lack of oxygen slows down the 

decomposition of organic matters. Sometimes the remains of animals can be found with hunting marks from 

arrows or spears. Well-preserved human bodies have been excavated from moors and bogs. The anaerobic 

conditions preserve the bodies very well and several thousands of years later they are excavated with skin, 

flesh and clothes still present. Wooden constructions, such as poles for bridges, boats and wooden tools may 

also be preserved, giving us valuable information on the level of technology at that time. 

Past farming practices can also be recognized in the soil profile, particularly in Anthrosols. For example, in 

northwest Europe, notably in the Netherlands and Germany, a human-made soil type, known as plaggen 

soil, has developed as a result of a specialized agricultural system. On the strongly leached, acid sandy outwash 

plains and moraines, Podzols developed underneath a vegetation cover of heather. Farmers used the heather 

and the uppermost level of the soil as bedding in the stables. The droppings from the animals, mixed with the 

bedding, were later used as manure on the nearby fields, slowly building up a thick soil layer rich in organic 

matter and high in nutrients and with a good soil water retention. These fields provide a relatively high and 

stable crop production compared to the surrounding land (European Soil Bureau Network, 2005). 

In the Amazon basin, the Terra Preta soil owes its name to its very high charcoal content. It was created 

by the addition of a mixture of charcoal, bone and manure to the otherwise relatively infertile Amazonian 

soil. Terra Preta soils were created by indigenous peoples in the pre-Columbian era between 450 BC and 

AD 950 (Sombroek et al., 2002). Technosols are modern examples of soils that store artefacts or are strongly 

influenced by (modern) humankind. They include soils from wastes (landfills, sludge, cinders, mine spoils and 

ashes), pavements with their underlying unconsolidated materials, and constructed soils in human-made 

materials (FAO, 2014). 

Soils provide aesthetic and recreational value through the landscape, particularly in Globally Important 

Agricultural Heritage Systems (Koohafkan and Altieri, 2011). They have also been used as an aesthetic approach 

to raise soil awareness in contemporary art (Toland and Wessolek, 2014). Churchman and Landa (2014) provide 

a comprehensive treatment of the topic. 

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8 | Governance and policy 

responses to soil change 


This chapter provides an overview of policy and governance responses to soil change. While most attention 

is given to issues at the global, regional and national levels, it is emphasized that effective responses nearly 

always have a basis in local action by individual land managers. Indeed, understanding the interconnectedness 

and the consequences of actions at each level is central to effective governance and policy. 

This book, and in particular the regional assessments of soil change (Chapters 9 to 16), demonstrate that 

at the global scale there is a qualitative appreciation of the pressures on soil resources but limited consistent 

evidence on their condition and trajectories of change. These assessments reveal that some of the world’s soil 

management challenges are immediate, obvious and serious – they arise partly because of the nature of soils 

in different regions and their associated history of land management. Other problems are more subtle but 

equally important in the long term – they require vigilance and a sustained policy response over decades. At 

present, few countries have effective policies to deal with these problems. In short, the world’s soils need to 

support at least a 70 percent increase in food production by 2050 (FAO, 2011) but there are some fundamental 

uncertainties. For example: 

Is there enough arable land with suitable soils to feed the world in coming decades? 

Are soil constraints partly responsible for the apparent yield plateau for major crops? 

Status of the World’s Soil Resources | Main Report Governance and policy responses to soil change 

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Can changes to soil management have a significant impact on the seemingly unsustainable global demand 

for nutrients? 

Can changes to soil management have a significant impact on atmospheric concentrations of greenhouse 

gases without jeopardising other functions such as food and fibre production? 

Will the extent and rate of soil degradation threaten food security and the provision of ecosystem services 

in the coming decades? 

Can water-use efficiency be improved through better soil management in key regions facing water scarcity? 

How will climate change interact with the distribution of soils to produce new patterns of land use? 

A comprehensive global view is needed to respond to these questions. A comprehensive view is also needed 

to deal with the trans-national aspects of food security and soil degradation. Through trade, most urbanised 

people are protected from local resource depletion. The area of land and water used to support a global citizen 

is scattered all over the planet. As a consequence, soil degradation and loss of production are not just local or 

national issues – they are genuinely international. 

The consideration of soil in policy formulation has been weak in most parts of the world. Reasons for this 

weakness include the following. 

Lack of ready access to the evidence needed for policy action. 

The challenge of dealing with a natural resource that is often privately owned but is at the same time an 

important public good. 

The long-time scales involved in soil change – some of the most important changes take place over decades 

and they can be difficult to detect. As a result, communities and institutions may not respond until critical and 

irreversible thresholds have been exceeded. 

Perhaps even more significant for policy makers is the disconnection between our increasingly urbanized 

human societies and the soil. The task of developing effective policies to ensure sustainable soil management is 

neither simple to articulate nor easy to implement. This is true regardless of a country’s stage of development, 

its natural endowment of soil resources, or the threats to its soil function. 

8.2 | Soils as part of global natural resources management 

In setting the stage, it is useful to examine the major drivers, pressures and institutional responses to land 

use and then set these within the broader international sustainable development agenda (see Table 8.1). 

8.2.1 | Historical context 

The ‘Great Dust Bowl’ of the 1930s in the United States of America was pivotal because it triggered 

widespread public concern about land use, degradation and the need for sustainable management. Severe 

wind erosion resulted from the opening up of vast areas for cereal production through mechanisation, with 

associated loss of protective vegetation cover. In response, the Soil Conservation Service of the USDA was 

established in 1935. This served as a model for many other countries facing similar issues (Young, 1994). In 1937, 

the United States President Franklin D. Roosevelt famously stated ‘The nation that destroys its soil destroys 

itself.’ This is perhaps the most succinct and sharpest challenge for policy makers and it remains an all-too-real 


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contemporary challenge for policy makers in many countries. 

After World War Two, many countries experienced food shortages and governments responded by increasing 

their investments in agricultural research. Understandably, most of this research focussed on increasing crop 

yields and food production. During this period there was also a large investment in soil and land resource 

surveys, particularly in Africa and Asia. In the following decades, soil science was strongly supported, with 

diverse institutional responses emerging in different regions. In some countries, there was close integration 

with other aspects of natural resource management while in others, separate soil agencies were established. 

The FAO played an important role in developing influential technical standards (e.g. FAO, 1976) and supporting 

within-country programs that aimed to establish sustainable soil and land management. The production of 

the FAO-UNESCO (1980) Soil Map of the World was a landmark achievement. 

The success of the Green Revolution (Borlaug, 2000) along with large increases in crop yields in North 

America and Europe eventually led to less investment by public agencies in agricultural science and related 

activities. The emphasis shifted to environmental issues, a transition which occurred during the 1970s and 

1980s, particularly in developed western countries. During the 1990s and 2000s, disinvestment in soil science 

was widespread and many soil departments in universities or governments were either closed or incorporated 

into natural resource or environmental units. The UN commitment to soil resources through the FAO and 

related agencies was also scaled back dramatically. 

The food price rises in 2007 and 2008 shocked many policymakers out of the belief that stable or declining 

food prices and assured supplies could be taken for granted (Beddington et al., 2012). This period also marked 

the start of a critical re-examination of the capacity of the world’s soil resources to support sustainable 

agriculture, assist with climate regulation, and safeguard ecosystem services and biodiversity. Before exploring 

this topic in more detail, it is useful to review some of the key global agreements relating to soils that emerged 

from the 1980s onwards. 

8.2.2 | Global agreements relating to soils 

In 1982, the FAO adopted the World Soil Charter and UNEP published the World Soils Policy (FAO, 1982; 

UNEP, 1982). It has been difficult to assess the practical impact of these initiatives. Nevertheless, the principles 

and definitions provided useful guidance for national governments that pursued actions on sustainable soil 

management. 

The first United Nations Conference on Environment and Development (UNCED, 1992, also known as 

the ‘Earth Summit’) launched the global environmental agenda (Table 8.1). The UN Convention to Combat 

Desertification (UNCCD) addressed issues of desertification, land degradation and drought; the UN 

Framework to Combat Climate Change (UNFCCC) was to tackle climate change; and the Convention on 

Biological Diversity (CBD) dealt with the challenges of biodiversity conservation and sustainable use (CBD). 

Supported by the Global Environment Facility (GEF), these conventions have raised awareness and mobilised 

increased efforts by countries and partners to generate global environmental benefits. These conventions also 

cover, albeit with less prominence, issues of soil conservation, sustainable land management and land use 

change, taking into account human as well as ecological perspectives (Hurni et al., 2006). 

The ecosystem approach promoted by the CBD between 1998 and 2004 (CBD, 2014), recognised that 

human management is central to biodiversity conservation and sustainable use. This ecosystem approach 

was further developed in the Millennium Ecosystem Assessment of 2005. This paved the way amongst 

international agencies and donor funds for more integrated ecosystem approaches in agriculture. These 

approaches emphasized the need for sectoral integration, with increased attention given to the benefits of 

mixed agroforestry and agro-silvo-pastoral systems. Similar approaches had already been developed in many 

countries. Globally, there was a trend towards the use of incentive measures to encourage land users to 

adopt sustainable practices which not only enhance production but also maintain biodiversity and ecosystem 

services (FAO, 2007; MA, 2005; UNEP, 2004). Soils came to be seen in relation to the services they provide for 

human well-being and poverty reduction. However, compared to other functions, soil-related matters did not 


225

feature prominently in policy or programmes. 

Following the food crisis in 2008, policy makers at the international level began to appreciate that soils were 

finite and an important factor that had to be considered in the debate on food security. Within the framework 

of UNCCD’s ‘Zero Net Land Degradation’, discussions were initiated about the need for quantitative targets 

and indicators to measure soil degradation (UNCCD, 2012). Concerns over food insecurity, water scarcity, 

climate change and increasing pressures on limited land and water resources led to much greater dialogue, 

advocacy and partnerships supporting integrated approaches to this complex set of issues (Beddington et al., 

2012; Steffen et al., 2015). 

The UN Conference on Sustainable Development (Rio+20), took place in June 2012, two decades after the 

Earth Summit. In the resulting document, The Future We Want, the international community agreed on the 

need to achieve a land degradation neutral world in the context of sustainable development (UN, 2012). The 

conference also initiated the process of developing universal Sustainable Development Goals (SDGs). 

All of the above developments relating to soils and land degradation are framed by the broader issue of 

climate change. Again, there is a long institutional history but it is useful to start with the establishment of 

the Intergovernmental Panel on Climate Change (IPCC) in 1988 by the UN Environment Programme (UNEP) 

and the World Meteorological Organization (WMO). The IPCC provides the world with scientific and technical 

information on climate change and its socio-economic impacts. The next major development was adoption 

of the Kyoto Protocol in 1997 by the UNFCCC. The Protocol, which entered into force in 2005, committed 

industrialized countries to stabilize greenhouse gas emissions, in particular carbon dioxide (CO 2 

). The Protocol 

started as a non-binding agreement but later progressed to legally binding agreements on emission reduction 

targets. The Protocol is of great importance for soils and land management because soils are important 

carbon sinks. The Protocol recognized opportunities for better management of carbon stores and for the 

enhancement of carbon sequestration in forestry and agriculture. There was thus clear recognition that soil 

management can be a vehicle to achieve climate goals – and conversely, that soils can be managed to avoid 

the loss of carbon through land degradation. Because of the climate system’s sensitivity to soil processes, soilrelated 

issues are set to attract increasing attention in future climate agreements. 

In recent years, FAO and its member countries have made significant progress in supporting strategies 

and policies to improve global governance of soil resources. In order to meet the need for a multilateral 

agreement focusing specifically on soil challenges, and to advocate for sustainable soil and land management 

at global level, the Global Soil Partnership1 (GSP) was proposed by FAO and the EU and then established in 

September 2011. The GSP strives to raise awareness among decision makers on the role of soil resources in 

relation to food security, climate change, and the provision of ecosystem services (Montanarella and Vargas, 

2012). Technical and scientific guidance is provided by the Intergovernmental Technical Panel on Soils (ITPS). 

The ITPS complements related scientific advisory panels including the Intergovernmental Panel on Climate 

Change (IPCC), the Intergovernmental Panel on Biodiversity and Ecosystem Services (IPBES), and the UNCCD’s 

Science-Policy Interface (SPI). 

The ITPS has been key to the development of the Plans of Action for the five pillars of the Global Soil 

Partnership (Table 8.2). It has also been engaged in the development of the Sustainable Development Goals 

and the initiation of formal reporting mechanisms, including the present book. An indication of the emerging 

priority accorded to soils and a measure of the impact of the GSP was the declaration by the United Nations 

General Assembly of 2015 as the International Year of Soils. 

1 www.fao.org/globalsoilpartnership 


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Year 

1982 FAO World Soil Charter 

1988 Intergovernmental Panel on Climate Change (IPCC) 

UN Conference on Environment and Development 

Rio Declaration 

Agenda 21 

1992 

Global Environmental Facility 

UN Convention to Combat Desertification (UNCCD) 

UN Framework to Combat Climate Change (UNFCCC) 

Convention on Biological Diversity (CBD) 

1997 Kyoto Protocol 

2000 Millennium Development Goals (MDGs) 

2005 Millennium Ecosystem Assessment 

2008 UNCCD’s Zero Net Land Degradation 

2011 Global Soil Partnership initiated (FAO/EU) 

2012 

Rio+20 

Sustainable Development Goals (SDGs) and Post-2015 Development Agenda 

Intergovernmental Technical Panel on Soils (ITPS) of the GSP 

2013 

Updated FAO World Soil Charter 

Land and Soils integrated in the Open Working Group of the Sustainable Development 

Goals 

Regional Soil Partnerships of the GSP 

2015 International Year of Soils declared by the UN General Assembly 

Table 8.1 Recent Milestones in soil governance and sustainable development 

Pillar No. 

1 

2 

3 

4 

5 

Action 

Promote sustainable management of soil resources for soil protection, 

conservation and sustainable productivity 

Encourage investment, technical cooperation, policy, education awareness 

and extension in soil 

Promote targeted soil research and development focusing on identified gaps and priorities 

and on synergies with related productive, environmental and social development actions 

Enhance the quantity and quality of soil data and information: data collection, analysis, 

validation, reporting, monitoring and integration with other disciplines 

Harmonize methods, measurements and indicators for the sustainable management 

and protection of soil resources 

Table 8.2 The 5 Pillars of Action of the Global Soil Partnership. 


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8.3 | National and regional soil policies 

8.3.1 | Sustainable soil management – criteria and supporting practices 

International agreements on soil and land resources are helpful but they are all to no avail unless there are 

complementary policies and coordinated activities at regional, national, district and local levels. Appropriate 

and effective policies need to reflect the local context in terms of the natural resource issues, culturally 

acceptability and economic feasibility. However, a unifying scientific narrative is also needed. In broad 

terms, the criteria for determining whether a landscape is functioning effectively and whether soils are being 

managed sustainably are as follows. 

Leakage of nutrients is low. 

Biological production is high relative to the potential limits set by climate and water availability. 

Levels of biodiversity within and above the soil are relatively high. 

Rainfall is efficiently captured and held within the root zone. 

Rates of soil erosion and deposition are low, with only small quantities being transferred out of the system. 

Contaminants are not introduced into the landscape and existing contaminants are not concentrated to 

levels that cause harm. 

Systems for producing food and fibre for human consumption do not rely on large net inputs of energy 

Net emissions of Greenhouse Gases are zero or less. 

We can manage what we can measure, so the task is to ensure that the above criteria can be measured 

against locally appropriate benchmarks. Without this information, policy makers and land managers do 

not have indicators of whether they are moving towards sustainability or going backwards. Policy makers 

also require an appreciation of how soil and land management practices can be applied to achieve desired 

outcomes. Regardless of the level of mechanization and technological sophistication, farming practices in 

general need to (FAO, 2013): 

Minimize soil disturbance by avoiding mechanical tillage in order to maintain soil organic matter, soil 

structure and overall soil function. 

Enhance and maintain a protective organic cover on the soil surface, using cover crops and crop residues, in 

order to protect the soil surface, conserve water and nutrients and promote soil biological activity. 

Cultivate a wide range of plant species – both annuals and perennials – in associations, sequences and 

rotations that include trees, shrubs, pastures and crops, in order to enhance crop nutrition and improve 

system resilience. 

Use well-adapted varieties with resistance to biotic and abiotic stresses and with improved nutritional 

quality, and to plant them at an appropriate time, seedling age and spacing. 

Enhance crop nutrition and soil function through crop rotations and judicious use of organic and inorganic 

fertilizer. 


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Ensure integrated management of pests, diseases and weeds using appropriate practices, biodiversity and 

selective, low-risk pesticides when needed. 

Manage water efficiently. 

Control machines and field traffic to avoid soil compaction. 

Thousands of different soil and land management practices have been developed around the world in 

response to local biophysical, social and cultural settings (e.g. WOCAT, 2007). Most cultures have deep 

connections with the land, and soil is venerated in diverse ways (Churchman and Land, 2014). In many 

regions, traditional knowledge still plays an important role in determining land management. However, most 

traditional systems have been disrupted or modified for a wide range of reasons. The two most common 

reasons have been the loss of access to land (e.g. invasion and displacement; increasing population densities 

causing shorter fallow periods on smaller areas; loss of access to grazing lands) and the arrival of new 

technologies. 

8.3.2 | Education about soil and land use 

Regardless of the culture or landscape setting, knowledge of soil and land resources is the foundation for 

achieving sustainable soil management (Dalal-Clayton and Dent, 2001). Spreading knowledge about soils 

requires formal education, preferably at all levels of schooling. Some countries are developing comprehensive 

and imaginative curricula that use an understanding of soils as a basis for teaching a wide range of cultural, 

social, scientific and economic subjects. At a more advanced level, training is needed in a range of soil science 

sub-disciplines (e.g. soil physics, soil chemistry, soil biology and pedology), Training in soil science needs to be 

linked to related disciplines including geology, ecology, forestry, agronomy, hydrology and other environmental 

sciences. Mechanisms for outreach, vocational training and extension are also needed. 

Policy makers need to ensure that education systems provide sufficient understanding and training for 

a nation to achieve sustainable soil management. In particular, farmers and others directly involved in soil 

management require sufficient knowledge to manage their soils profitably and sustainably. 

8.3.3 | Soil research, development and extension 

The second key area where policy makers have responsibility is in relation to research, development 

and extension. The pioneering work of the Soil Conservation Service in the United States and the technical 

innovations of the Green Revolution are two examples that demonstrate the power of agricultural science and 

technology. The Green Revolution also highlights how trade-offs are required when there is a focus on a single 

ecosystem service (food production) at the expense of others (e.g. water quality). Contemporary science policy 

often focuses on impact and public benefit. In this regard, soil research is often considered simply as a means 

to an end. Although soil science is vital to several important ends, notably agriculture, environment, water 

management and climate change, it is often overlooked in priority setting exercises. More formal recognition 

of soil resources as a cross-cutting issue in science policy is necessary to ensure it receives sufficient support. 

The recent Australian initiative to achieve a more integrated view of soil research, development and extension 

is instructive in this regard (Australian Government, 2014). 

8.3.4 | Private benefits, public goods and payments for ecosystem services 

The amount of regulation on land use and management varies substantially between countries depending 

largely on the degree of government intervention. Effective regulations on land use and management require 

a good information base for setting critical limits, implementing various zoning schemes and monitoring 


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compliance. In practice, regulating soil management practices (e.g. application of manure, moderating or 

increasing fertilizer use, control of dryland salinity) and implementing zoning systems (e.g. to protect the best 

agricultural soils) involves complex technical, institutional and policy challenges. 

Countries that rely less on regulation often opt for incentive schemes to achieve outcomes. Incentives can 

range from subsidy systems (e.g. for fertilizer in poor countries or for equipment for conservation tillage in 

developed countries) through to various forms of certification for the adoption of specified soil management 

practices (e.g. organic farming). Some of these systems have strong economic drivers because they are 

mandatory for market access (e.g. participation in supply chains to supermarkets). 

Implementing effective policies requires organized systems for monitoring soil conditions and an 

understanding of the relationship between soils and land management. Without this basic information, policy 

makers have no way of knowing whether regulations and incentive schemes are achieving the desired result. 

8.3.5 | Intergenerational equity 

Ensuring intergenerational equity is becoming more difficult as human pressures on soil resources reach 

critical limits. Most traditional cultures and systems of family farming have strong cultural norms that ensure 

tribal lands or family farms are passed to the next generation in the same or better condition than when they 

were inherited. However, dramatic changes to land management associated with intensive agriculture, the 

adoption of Green Revolution technologies, and intensification of land use more generally, are having a major 

impact on soil resources. The area of arable land per capita is decreasing sharply (0.45 ha in 1961, 0.25 ha 

in 2000 and a forecast of 0.19 ha in 2050). Future generations will inherit a radically modified land and soil 

resource. 

Many countries have sophisticated reporting systems for assessing issues relating to intergenerational 

equity (e.g. long-term forecasts to determine the viability of pension and health systems; decadal plans for 

critical infrastructure). Scenario analysis and futures forecasting are essential to national preparedness and 

long-term sustainability. There is now an imperative for policy makers to assess the current trends in soil 

condition and natural resource scarcity summarised in this book and to factor in the consequences to scenario 

analysis and futures forecasting. 

8.3.6 | Land degradation and conflict 

Land degradation and resource scarcity can play a role in the rise of conflicts, but these conflicts are rarely 

purely resource driven. Where tensions about access and use of natural resources do exist, they depend on 

a variety of factors – the outcomes of which may sometimes cascade from tension into violent conflict, but 

certainly not always. More often than not, natural resource degradation is a result of conflict rather than a 

cause. The existence of land degradation can also lead people to seek cooperative solutions. Policy makers 

and others involved in land management can not only act to resolve resource conflicts but also help to prevent 

them and to find peaceful mutually acceptable solutions (Frerks et al., 2014; Bernauer, Böhmelt and Koubi, 

2012). 


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8.4 | Regional soil policies 

8.4.1 | Africa 

Africa has a diverse range of soils and land use systems. However, very large areas, particularly in West 

Africa, are infertile or of low fertility, and unsustainable systems of land use are widespread. A leading cause 

of low fertility is nutrient depletion (Smaling, 1993; Stoorvogel, Smaling and Janssen, 1993). This is considered 

to be the chief biophysical factor limiting small-scale farm production (Drechsel, Giordano and Gyiele, 2004) 

although other factors including limited organic matter and erosion are significant as well (Bossio, Geheb 

and Critchley, 2010). Mounting concern over these issues contributed to the creation of the New Partnership 

for Africa’s Development (NEPAD). This is a vision and policy framework produced by the African Union (AU) 

that aims to provide member countries with guidance over their development agenda. Within NEPAD, the 

Comprehensive Africa Agriculture Development Programme (CAADP) sets out an agenda targeting annual 

growth of 6 percent in agricultural production. The Abuja Declaration on Fertilizers, agreed in 2006, laid 

out the vision for an African Green Revolution. Central to this was the aim of increasing the level of fertilizer 

application from 8 kg ha -1 to 50 kg ha -1 . However, only slow progress has been made in implementing this 

agenda at regional and country level (NEPAD, 2012). 

Food policy and agricultural development in Africa pose challenges beyond the scope of this book. However, 

there are some promising developments even for countries facing the most daunting difficulties owing to rapid 

population growth, very low incomes, weathered and infertile landscapes, low levels of literacy, vulnerability 

to climate variability and change, disease and significant potential for social unrest. Two of these promising 

developments have been supported by the Bill and Melinda Gates Foundation. 

First is the AGRA Soil Health Programme which aims to increase income and food security by promoting 

the wide adoption of integrated soil fertility management (ISFM) practices among smallholder farmers and 

creating an enabling environment for wide adoption of these improved practices across sub-Saharan Africa. 

The objective is to improve supply and access to appropriate fertilizers, as well as access to knowledge on IFSM 

for over four million smallholders and to strengthen extension and advisory capacity. The Programme also 

seeks to influence national policy in favour of investment in fertilizer and ISFM. Some 1.8 million smallholders 

are reported to be using ISFM, including fertilizer micro-dosing, manure and legumes in crop rotations, with 

yields in the Sahel up three to fourfold in good seasons. 

The second promising initiative is the AgWaterSolutions Project. The project concept builds on the existence 

of sizable untapped groundwater systems in the region and on the recent availability of small affordable 

motorized water pumps. The project promotes small-scale distributed irrigation systems that rely primarily 

on groundwater. In these systems, the access point for water, the distribution system and the irrigated crop 

all occur at or near the same location. These systems are typically privately owned and managed by individuals 

or small groups. The potential in countries such as Burkina Faso is large. This initiative is helping to shift the 

attention of policy makers and planners away from large scale irrigation developments. 

There are many other significant soil policy issues facing the region. Examples include: the costs and benefits 

of subsidy schemes for fertilizers; the growing pressure on land resources and the consequent shortening of 

fallow periods; the challenge of making inputs affordable and ensuring market access in areas where poverty 

is prevalent; and addressing urban and peri-urban planning so that more intensive and safe food production 

systems can develop in and around the rapidly growing African cities. 


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8.4.2 | Asia 

In regions of rapid development in Asia, urbanization, industrialization and intensive land use lead to 

unbalanced use of agro-chemicals, poor waste management and acid deposition caused by urban air 

pollution. These factors have contributed to increasing soil contamination and acidification. In China, for 

example, a soil pollution survey found that 6.4 million square kilometres of arable land are contaminated and 

that this represents an alarming threat to human health (Yue, 2014). In consequence, China’s Environmental 

Protection Law was revised and strengthened in 2014. However, in China and all across the region the greatest 

environmental challenges arise from the gap between legislation and implementation (Mu et al., 2014). 

In recent years government policy responses across the region have encouraged improved land use practices 

that increased tree cover for carbon sequestration. Carbon-financing schemes have been implemented. 

However, government policies have been less effective in dealing with the issue of foreign investment in 

agricultural land. In some countries, foreign companies have begun a variety of contractual arrangements 

with local farmers, resulting in some cases in the loss of land for smallholders (Fox et al., 2011). 

8.4.3 | Europe 

Europe has well-established and strong formal governance mechanisms to address environmental issues 

at regional, national and sub-national levels. European Union (EU) environmental policies are agreed at 

central level but legislated and implemented at the national level. However, the experience with soils policy 

has been more complex and only a handful of member states have specific legislation on soil protection. With 

the objective of protecting soils across Europe, the European Commission adopted a Soil Thematic Strategy 

in 2006 which consists of a communication, a proposal for a framework directive (under European Union 

legislation) and an impact assessment (EC, 2006). The proposal for a Soil Framework Directive would require 

member states to adopt a systematic approach to identifying and combating soil degradation. However, this 

could not be agreed by the required majority in the European Council and the draft Directive was consequently 

withdrawn by the European Commission at the end of 2014. The failure to adopt the directive was largely 

due to concerns about subsidiarity, with some member states maintaining that soil was not a matter to be 

negotiated at the European level. Others felt that the cost of the directive would be too high, and that the 

burden of implementation would be too heavy. However, the Seventh EU Environment Action Plan, which 

entered into force in 2014, recognises the severe challenge of soil degradation. It provides that by 2020 

land in the EU should be managed sustainably, soil should be adequately protected, and the remediation 

of contaminated sites should be well underway. Furthermore it commits the EU and its member states to 

increasing efforts to reduce soil erosion, to increase soil organic matter and to remediate contaminated sites 

(EC, 2013). 

8.4.4 | Eurasia 

Eurasian countries have well-developed environmental policies and regulations. However following the 

break-up of the Soviet Union, the system of environmental monitoring and conservation collapsed and has 

only recently been partially restored. Countries all across the region have maintained and even improved 

environmental and soil conservation legislation in recent years, but in most countries the mechanisms for 

quality control and environmental monitoring have been weakened. For example, only Belarus and Uzbekistan 

maintain their soil survey institutes, and even there soil monitoring has been discontinued. 

Ukraine, Russia and Kazakhstan are the countries with the largest under- or unused agricultural lands in 

the world. The World Bank (2011) states that these countries have the capacity to meet the growing global 

demand for food. In Russia in 2002 the area of abandoned land reached 70 million ha. Since then there has 

been a slow decrease in the area of unused land (Nefedova, 2013). However, it should be noted that most 

land abandonment occurred in badlands, wetlands, steep slopes and areas with an unfavourable climate, 


232

while in areas with fertile soils the investment in land management increased. In countries such as Ukraine 

and Georgia, where land tenure legislation allows land ownership by non-residents, foreign capital is being 

invested in farmland. Non-transparent land grabs on a large scale are expected to increase, and might have far 

reaching consequences for the livelihoods of the rural population (Visser and Spoor, 2010). 

8.4.5 | Latin America and the Caribbean (LAC) 

This region is one of the richest in the world in terms of natural resources. However, rapid exploitation and 

export of these resources (minerals, gas, forests, and pastures) is occurring with associated dramatic land 

use changes and widespread land degradation. Nonetheless, some countries in the region have developed 

and implemented good policies and approaches to mitigate land degradation. These policies, implemented 

at national and sub-national levels, are good practice examples that could be replicated in other countries in 

the region (UNEP, 2012). 

Uruguay provides a good example of soil and land conservation policies: here the soil conservation policy 

was designed by the Ministry of Livestock, Agriculture and Fisheries (MGAP) within a programme promoting 

agricultural intensification, with the objective of implementing a sustainable intensification model. Under 

this policy, crop producers must submit soil management plans and state the rotation sequence on each plot. 

They must stay within the maximum tolerable soil erosion amount based on local soil characteristics (Hill, 

Mondelli and Carrazzone, 2014). 

Another example is Cuba’s National Environmental Strategy of 2011/2015 which characterizes soil 

degradation as one of the fundamental environmental challenges in the country. The Cuban government 

has also implemented action plans to fight desertification and, since 2001, has undertaken programmes for 

soil conservation (CITMA, 2011). Brazil’s Forest Code was updated in 2012: it establishes general standards 

for protection of forests and other native resources, including soil and water resources. The Forest Code 

also integrates legal and economic incentives to promote sustainable production activities. However, closer 

analysis of the updated Forest Code suggests that it may in fact allow more deforestation than the previous 

version, in response to the demands of agricultural intensification (Soares-Filho et al., 2014). 

8.4.6 | The Near East and North Africa (NENA) 

This region is considered as the most water scarce and arid region in the world. Moreover, given the 

scarcity of land and water resources, this region is particularly vulnerable to the impacts of climate change, 

increasing drought, declining soil fertility and consequently declining agricultural production (Wingkvist and 

Drakenberg, 2010; Drine, 2011). There are government programs to improve land management in several 

countries, especially countries that are party to international agreements and are in receipt of donor support. 

Most actions promoting sustainable land management have been to combat desertification under the 

framework of the UNCCD (UNCCD, 2012). 

Despite significant improvements in the region in tackling the root cause of land degradation, there are 

still challenges in enforcing environmental regulations and implementing environmental conservation 

policies. The main implementation constraints are: the weakness of institutions at all levels; the difficulty of 

coordinating action across sectors, themes, donors and stakeholders; the lack of participation of the local 

communities; and tenure insecurity. 

MENARID (Integrated Natural Resources Management in the Middle East and North Africa) is a 

partnership working for improvement of the governance of natural resources, including water. MENARID 

supports restoration of natural resources. In particular, the programme aims to improve the livelihoods of 

target communities through the restoration of degraded natural resources, including land and soils. It offers 

a platform for coordination between stakeholders and information sharing in the countries (ICARDA, 2013). 


233

The NENA region is endowed with oil and gas reserves. In areas of rapid urbanization and oil production, 

soil pollution and soil sealing are associated challenges. Parts of the region are extremely sensitive to political 

conflicts, and peace and post-conflict are the main focus. Land degradation issues become more pronounced, 

but inevitably they have to take second place to other concerns. 

8.4.7 | North America 

In the United States, federal policies favour market-based instruments within an overall environmental 

governance framework, and these instruments have superseded traditional regulatory instruments. Land use 

is a priority issue on the political agenda, due to its contribution to GDP through forestry and agriculture. 

Governments diminish environmental impacts by paying land managers to implement sustainable land 

management practices and soil conservation. Taxes and incentives encourage land and farmland preservation 

programmes through payment for ecosystem services (UNEP, 2012). 

The United States Conservation Reserve Program (CRP) pays farmers to remove land from agricultural 

production in order to prevent soil erosion and improve ecosystem functions. This set-aside generates 

economic benefits of around US$1.3 billion per year (Hellerstein, 2010). However, high prices have made 

agriculture more profitable and the rates of payment from CRP have not risen so fast. The amount of land 

enrolled in the programme is therefore expected to decline (Wu and Weber, 2012). The Environmental Quality 

Incentives Program and the Observation Security Program of 2002 are other programmes that reward 

farmers for applying sustainable land management practices. It has been estimated that soil erosion could be 

reduced by 17 percent, saving around 36 million tonnes of soil annually. Valued at US$2 per tonne, the cost of 

conservation would thus be US$34 million annually, compared to the cost of restoring the soils, estimated at 

up to US$332 million (Hellerstein, 2010). 

In Canada land-use planning is a provincial responsibility and legislation differs widely among provinces. 

British Columbia has a long-standing Agricultural Land Reserve Program that prohibits development on 

approximately 4.7 million ha of agricultural land throughout the province. In the early 2000s Ontario created 

a Greenbelt that protects 0.7 million ha of agricultural and natural lands in the most populated region 

surrounding Toronto. Generally in Canada the implementation of Payment for Ecosystem Services needs still 

to be complemented with land use planning frameworks in order to become more effective at all levels of 

government (Calbick, Day and Gunton, 2003). 

8.4.8 | Southwest Pacific 

The scale of land degradation across the countries of the Southwest Pacific has given rise to a range of 

significant policy responses all with a strong emphasis on participative engagement and local action. Perhaps 

most significant has been the rise of the Landcare Movement in Australia. It began with an unlikely alliance 

between traditional opponents (conservationists and farmers) and grew into a movement with thousands of 

groups in Australia and in other countries. The activities of Landcare Groups transformed many landscapes 

with large areas being revegetated and restored. Youl et al. (2006) provide a good outline of the history and 

factors that were important for success. They conclude that the strength of Australian Landcare is that 

community groups and networks, with government and corporate support, conceive their own visions and 

set goals for local and regional environmental action. Working from the ground up to achieve these goals 

creates freedom and flexibility, giving communities a great sense of purpose. 

The Secretariat of the Pacific Community (SPC) is a regional intergovernmental organisation whose 

membership includes both nations and territories in the Pacific Ocean and their metropolitan powers. The 

Land Resources Division assists the Pacific Community to improve food, nutritional and income security and 

sustainable management and development of land, agriculture and forestry resources. 


234

In New Zealand, there are few regulatory instruments directly related to soil. Where they exist, they focus on 

soil conservation. However there is an increasing number of national policy instruments that legislate against 

the impacts of unwise soil use. The Resource Management Act is given effect at regional level and regulates 

activities not outcomes (through regional policy statements, plans and resource consents). These regulatory 

instruments typically focus on ensuring soil intactness. However, new initiatives are increasingly looking at 

the consenting of land use according to soil capability. New Zealand has also used non-regulatory approaches 

to achieve good soil management. These approaches include direct payments, support to the development of 

industry codes of practice, and certification schemes to ensure market access. 

8.5 | Information systems, accounting and forecasting 

The distribution and characteristics of the soils in any district or nation are neither obvious nor easy to 

monitor. As a consequence, understanding whether a land use is well-matched to the qualities of the soil 

requires some form of diagnostic system to identify the most appropriate form of management and to monitor 

how the soil is functioning. Important components of the diagnostic system necessary for sustainable land 

use and management are: 

an understanding of spatial variations in soil function (e.g. maps and spatial information) 

an ability to detect and interpret soil change with time (e.g. via monitoring sites, long-term experiments, 

environmental proxies) 

a capacity to forecast the likely state of soils under specified systems of land management and climates 

(e.g. through the use of simulation models) 

an understanding of the edaphic requirements of plants 

Preparation of this book was severely constrained by the lack of relevant information. Soil map coverages 

are variable and, in some regions, out-of-date. The capacity to monitor and forecast soil change is also 

rudimentary. All nations require coordinated soil information systems that parallel those that exist in many 

countries for economic data, weather and water resources. Action on soil information systems is enshrined 

in the World Soil Charter’s guidelines for action for governments (Sections VIII and IX) and international 

organizations (Sections I and II). However, creating appropriate institutional systems for soil information 

gathering and dissemination is challenging for the following reasons: 

All levels of government need reliable information on soil resources but often no single level of government 

or department has responsibility for collecting this information on behalf of other public sector agencies. 

Public and private interests in soil are large and overlapping – mechanisms for co-investment by public and 

private agencies are therefore needed. 

Market failure in relation to the supply and demand of soil information is a significant and widespread 

problem. Simply stated, beneficiaries of soil information do not usually pay for its collection and this reduces 

the pool of investment for new survey, monitoring and experimental programmes. 

Partly as a result of the above, soil-information gathering activities in many countries are currently funded 

through short-term government programmes, private companies or individuals or are produced in response to 

specific regulatory requirements. This piecemeal approach does not result in the kind of enduring, accessible 

and broadly applicable information systems that are needed to meet the requirements of stakeholders. 

The following sections outline some specific requirements that policy makers have of soil information 

systems. 


235

8.5.1 | Soil information for markets 

The various types of markets regulated by governments and other institutions need to be sufficiently 

informed to ensure economic efficiency and the desired allocation of resources. 

These markets include: 

traditional real-estate markets where information is needed on the capital value of soil resources (e.g. the 

nutrient status of a farm, presence of contaminants, and options for improved soil management) 

carbon trading schemes 

cap-and-trade systems for nutrient loading or other pollutants 

forecasting of within-season production of agricultural commodities 

insurance (e.g. crop insurance, disaster insurance, risk analysis of supply chains). 

Oversight and regulation of market activities is a central function of governments in most countries. A key 

responsibility for policy makers is to ensure the availability of reliable soil information. 

8.5.2 | Environmental accounting 

A closely related area where policy makers are starting to need better information is environmental 

accounting. Globally, national accounts of economic activity are recorded and indicators such as gross 

domestic product (GDP) are widely used in government and policy to assess economic activity and progress. 

However, indicators such as GDP measure mainly market-based transactions and are not a good indicator of 

welfare; GDP ignores social costs, environmental impacts and income inequality (Costanza et al., 2014). GDP 

also does not deduct the direct cost of the depletion of natural resources on national income nor does it take 

into account the impact that our resource extraction and use of nature has on the continued functioning of 

the Earth system for life support. 

In light of these limitations of the current national economic accounting system, the ecosystem services 

approach seeks to include nature in our accounting and acknowledge that it has value and its use is not simply 

free and limitless (Westman, 1977; Daily, 1997; Costanza et al., 1997; M 

A, 2005; Robinson et al., 2014). In this context, soils make an important contribution to the supply of 

ecosystem services (Daily et al., 1997; Wall, 2004; Robinson, Lebron and Vereecken, 2009; Dominati, Patterson 

and Mackay, 2010; Robinson et al., 2013). 

One proposal to address the deficiency of the current national accounts is to have a set of complementary 

accounts. Since the early 1990s, the international official statistics community has been developing such a set 

of accounts, named the System of Environmental Economic Accounting (SEEA). The over-arching objective 

of the SEEA approach is to develop an accounting structure that integrates environmental information 

with the standard national accounts and hence to mainstream environmental information in economic and 

development policy discussion. 

The SEEA accounts are presented in two volumes. First, the SEEA Central Framework (UN et al., 2014) 

which was adopted as an international statistical standard in 2012, and second, SEEA Experimental Ecosystem 

Accounting (UN et al., 2014) which was endorsed in 2013. The SEEA Central Framework deals with individual 

environmental assets (minerals, timber, fish, water, soil, etc.), the flows of mass and energy between the 

environment and the economy, and the space in which this occurs (Obst and Vardon, 2014). SEEA Experimental 

Ecosystem Accounting is focused on the function of ecosystems and the generation of ecosystem services 

which is dependent on ecosystem extent, condition and quality. 


236

The SEEA Central Framework identifies seven individual components of the environment as environmental 

assets; mineral and energy resources, land, soil resources, timber resources, aquatic resources, other biological 

resources (excluding timber and aquatic resources, for example, livestock, orchards, wild plants for medicine, 

wild animals that are hunted), and water resources. 

SEEA Experimental Ecosystem Accounting uses the same definition of environmental assets but rather 

than considering individual components as assets, it seeks to consider the way in which these components 

function jointly as ecosystems. To apply this logic it defines spatial areas, such as different vegetation habitats 

(forests, wetlands, agricultural land etc.) as ecosystem accounting units. In this approach soil is considered a 

component within a broader ecosystem rather than being considered as a distinct ecosystem. 

Soils form an important part of the Central Framework, being recognized as an environmental asset in their 

own right. An important distinction is made between land and soil resources. Land is considered in terms of 

space and location often referred to as Ricardian land (Daly and Farley, 2011). Soil resources are the volume 

of biologically active topsoil, and its composition in the form of nutrients, soil water and organic matter. 

The accounts are structured to recognize, and distinguish between, the use of an asset (e.g. soil volume 

and area within the asset accounts); and the use of the soil resource or elements of the soil resource (e.g. 

carbon, nutrients and soil moisture in the physical flow accounts). Fundamental to the accounting process 

is the measurement of change for both the environmental and ecosystem accounts, which is underpinned 

by the availability of good quality data (Obst, Edens and Hein, 2013). The major aspects of soil of interest for 

the environmental accounts are: the volume of soil moved or extracted; the area of soil under different land 

uses; carbon, nutrient and moisture stocks; and changes in these three aspects. Hence the understanding and 

quantification of soil change is central to environmental accounting (Robinson, 2015).There is still no agreed 

set of soil indicators, although soil carbon content is widely seen as being perhaps the main indicator. There 

is still much work to do to synthesize soil quality work into the SEEA framework for the creation of useful, 

informative accounts, and to encourage countries to adopt this unified approach. 

8.5.3 | Assessments of the soil resource 

It is essential to have some form of regular reporting on the rate and extent of soil change along with the 

likely consequences for society at local, national and global scales. Some countries now have various forms of 

audits and state-of-the-environment reports. However, most countries do not produce regular assessments 

showing where land management systems can operate sustainably within the constraints set by changing 

climate, weather and soils. These are necessary given the economic and environmental significance of soil 

resources. 

Regular reporting forces policy makers to impose an operational discipline on the management of soil 

information. Systems for collecting and analysing data can be progressively improved and a body of knowledge 

will be developed over several cycles of reporting. The assessments need to adopt a highly participative mode 

of engagement so that all stakeholders are represented and then empowered to make the necessary changes 

to land management. 

The World Soil Charter addresses this issue directly. It encourages governments to develop a national 

institutional framework for monitoring implementation of sustainable soil management and overall state 

of soil resources. International organizations are encouraged to facilitate the compilation and dissemination 

of authoritative reports on the state of the global soil resources and sustainable soil management protocols. 

This book is a sign that progress is being made in relation to regular assessment and reporting. Further 

progress will depend on successful implementation of Pillar Four of the Global Soil Partnership - Enhance 

the quantity and quality of soil data and information – and of Pillar Five - Harmonize methods, measurements and 

indicators for the sustainable management and protection of soil resources. 


237

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9 | Regional Assessment 

of Soil Changes in Africa South 

of the Sahara 

Regional Coordinator: Victor Chude (ITPS/Nigeria) 

Regional Lead Author: Ayoade Ogunkunle (Nigeria) 

Contributing Authors: Victor Chude (ITPS/Nigeria), Isaurinda Dos Santos (ITPS/Cape Verde), Tekalign Mamo 

(ITPS/Ethiopia), Garry Paterson (South Africa), Ndaye Soumare (Senegal), Liesl Wiese (South Africa), and 

Martin Yemefack (ITPS/Cameroon). 

Status of the World’s Soil Resources | Main Report Regional Assessment of Soil Changes 

242 

in Africa South of the Sahara


Land degradation in sub-Saharan Africa (SSA) is believed to be expanding at an alarming rate, accompanied 

by the lowest agriculture and livestock yields of any region in the world. While cereal production has increased 

marginally over the past two decades, more than 70 percent of this growth is due to area expansion rather 

than yield increases. The region also suffers from the world’s highest rate of deforestation, with some countries 

having lost more than 10 percent of their forest cover in the fiveyears up to 2009 (IFAD, 2009) and is most likely 

continuing at the same rate to this day. 

There is a growing and long-standing recognition among both policy-makers and soil specialists that soil 

degradation is one of the root causes of declining agricultural productivity in sub-Saharan Africa and that, 

unless the process of degradation is controlled, many parts of the continent will suffer increasingly from 

food insecurity (e.g. see Lal, 1990; UNEP, 1982). The consequences of allowing the productivity of Africa’s soil 

resources to continue on its present downward spiral will be severe, not only for the economies of individual 

countries, but for the welfare of the millions of rural households across the continent who are dependent on 

agriculture (FAO, 1999). 

Soil degradation is the decline in soil quality caused its improper use by humans, usually for agricultural, 

pastoral, industrial or urban purposes. Soil degradation may be exacerbated by climate change and 

encompasses physical, chemical and biological deterioration. Examples of soil degradation cited by Charman 

and Murphy (2005) are: loss of organic matter; decline in soil fertility; decline in structural condition; topsoil 

loss and erosion; adverse changes in salinity, acidity or alkalinity; and the effects of toxic chemicals, pollutants 

and excessive flooding. 

There is no consensus on the exact extent and severity of land degradation or its impacts in SSA as a whole 

(Reich et al., 2001; GEF, 2006). Lack of information and knowledge is considered to be one of the major 

obstacles for reducing land degradation, improving agricultural productivity, and facilitating the adoption of 

sustainable land management (SLM) among smallholder farmers (Liniger et al., 2011). The recent publication 

of the first Soil Atlas of Africa has provided a first comprehensive overview of the soil resources of Africa (Jones 

et al., 2013). 

Four continental-scale studies have assessed the extent of soil degradation in Africa. A literature review by 

Dregne (1990) of 33 countries found compelling evidence of serious land degradation in sub-regions of 13 

countries: Algeria, Ethiopia, Ghana, Kenya, Lesotho, Mali, Morocco, Nigeria, Swaziland, Tanzania, Tunisia, 

Uganda, and Zimbabwe. In another literature review, focused on drylands only, Dregne and Chou (1992) 

estimated that 73 percent of drylands were degraded and 51 percent severely degraded. They concluded that 

18 percent of irrigated lands, 61 percent of rainfed lands, and 74 percent of rangelands located in SSA drylands 

are degraded. 

The Global Assessment of Soil Degradation (GLASOD) expert survey found that 65 percent of soils on 

agricultural lands in Africa had become degraded since the middle of the twentieth century, as had 31 percent 

of permanent pastures, and 19 percent of woodlands and forests (Oldeman, Hakkeling and Sombroek, 1991). 

Serious degradation affected 19 percent of agricultural land. A high proportion (72 percent) of degraded land 

was in drylands. The most widespread cause of degradation was water erosion, followed by wind erosion, 

chemical degradation (three-quarters from nutrient loss, the rest from salinization), and physical degradation. 

In terms of causes of degradation, overgrazing was responsible for half of all degradation, followed by 

agricultural activities, deforestation, and overexploitation. 


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The Land Degradation Assessment in Drylands project (LADA) started in 2006 with the general purpose of 

creating the basis for informed policy advice on land degradation at global, national and local levels. This goal 

is being reached through the assessment of land degradation at different spatial and temporal scales in six 

countries and through the creation of a baseline at global level for future monitoring (FAO, 2010). Two of the 

six countries involved (Senegal and South Africa) are within SSA and national results are reported at the end 

of this chapter. 

Lal (1995) calculated continent-wide soil erosion rates from water using data from the mid to late 1980s, and 

then used these rates to compute cumulative soil erosion for 1970-90. The highest erosion rates occurred in the 

Maghreb region of Northwest Africa, the East African highlands, eastern Madagascar, and parts of Southern 

Africa. Excluding the 42.5 percent of arid lands and deserts with no measurable water erosion, Lal found that 

the land area affected by erosion fell into the following six classes of erosion hazard: none, 8 percent; slight, 49 

percent; low, 17 percent; moderate, 7 percent; high, 13 percent; and severe, 6 percent. 

Soils host the majority of the world’s biodiversity and healthy soils are essential to securing food and fibre 

production. Soils assure an adequate and clean water supply over the long term, as well as providing cultural 

functions. Ecosystem services provided by soils are integral to the carbon and water cycles. Major increases 

in agricultural production have been associated with different kinds of soil degradation, especially since 

the agricultural growth came in part from extensive clearing of new agricultural lands. Yet, even with this 

expansion, arable land per capita in Africa declined from just under 0.5 ha in 1950 to just under 0.3 ha in 

1990 (FAO, 1993). During this time period, yield increases on land already in production thus contributed far 

more to the total production. For example, more than 90 percent of the growth in developing country cereal 

production between 1961 and 1990 came from yield growth (World Bank, 1992). Agricultural expansion and 

yield growth at such a scale is inevitably associated with some degradation of soil resources. However, the 

type and extent of degradation vary in the different ecological/farming systems (IFPRI, 1999). 

9.2 | Stratification of the Region 

The region is diverse in terms of relief, climate, lithology, soils and agricultural systems. A combination of some 

of these have been used to stratify the region into agro-ecological zones (AEZs) (Fischer et al., 2002; Global 

HarvestChoice, 2010). Table 9.1 shows the AEZs into which the region has been grouped and some of their 

characteristics, while Figure 9.1 shows the distribution of the AEZs in the region. 

9.2.1 | Arid zone 

The arid zone occupies 36 percent of the land area of SSA, most of which is in West and East Africa. Rainfall 

is low and extremely variable in this zone. The annual rainfall of less than 500 mm, combined with high 

temperatures and consequent high rates of evapotranspiration, make this zone capable of sustaining plant 

life for less than 90 plant growth days (or length of growing season). 


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Figure 9.1 Agro-ecological zones in Africa South of the 

Sahara. Source: Otte and Chilonda, 2002. 

Zone 

Definition 

Rainfall 

range 

(mm) 

West 

Africa 

Central 

Africa 

Area (percent) 

East Southern 

Africa Africa 

Area 

of zone 

(percent) 

Arid 270 pgd 1500+ 10 59 2 7 19 

Highlandsb

The arid zone is mostly associated with sandy soils (Arenosols, Psamments) which are weakly differentiated 

and are often of aeolian origin. Water and air move freely through these soils, which are low in all nutrients. 

The accompanying vegetation consists of short annual grasses, legumes, scattered shrubs and trees. 

Mobile herds of sheep, goats, cattle and camels browse the herbage and shrubs, while farmers use most of the 

trees and shrubs for fuel. The low rainfall and its erratic distribution make cropping uncertain in most years. 

Owing to this unreliability, arable farming is mostly restricted to opportunistic cultivation of short-season 

millets, except in topographically favourable sites such as oases or irrigated areas. Opportunities for livestock 

development are limited but existing techniques could be improved upon, if not to increase productivity, then 

at least to sustain it. 

9.2.2 | Semi-arid zone 

The semi-arid zone receives 500 to 1000 mm of rainfall annually which can be capable of sustaining 90 

to 180 plant growing days. This zone occupies 18 percent of the land area of SSA. Semi-arid lands are found 

in all regions of SSA except central Africa. The low rainfall and the long dry season make the semi-arid zone a 

relatively healthy environment for humans and livestock. Arenosols (Psamments) and Cambisols (Inceptisols) 

are widespread and include coarse sandy soils, fine sands, and loamy sandy soils. Water retention is poor and 

nutrient contents, including N, P and S levels, are generally low. The permeability of the undisturbed soil is 

good, but algal skins contribute to the formation of surface crusts. The natural vegetation is an open low-tree 

grassland but this has been severely modified in many regions. 

The lower rainfall areas of this zone are used for grazing. Cropping and crop–livestock systems dominate the 

areas with higher rainfall where farmers commonly grow millet, sorghum, groundnut, maize and cowpeas. 

9.2.3 | Sub-humid zone 

The sub-humid zone occupies 22 percent of SSA, mainly in southern and central Africa. The zone receives 

1 000 to 1 500 mm of rain annually which can sustain plants for 180 to 270 plant growing days. Within the 

climatic definition, this is a very varied zone in terms of climate, soils and land use. Luvisols (Alfisols) and 

Cambisols (Inceptisols) occur widely, parent material is often strongly weathered, and the levels of mineral 

nutrients as well as the clay fraction are low. Cambisols (Inceptisols) have fewer constraints to plant production 

than the older, more weathered soils, since their high base status provides adequate Ca and eliminates 

constraints related to low pH levels. The fertility of many soils in this zone is low, especially due to leaching of 

NO 3 

–, accompanied by the loss of cations and P adsorbtion. In addition, structural stability in these soils can 

be poor, with crusting and hardening occurring when soils are dry. 

The natural vegetation is typically medium height or low woodland with understory shrubs and a ground 

cover of medium to tall, mainly perennial, grasses; Hyparrhenia spp. are common. 

Food and cash crops are grown, including cassava, yams, maize, fruits, vegetables, rice, millet, groundnut, 

cowpeas and cotton. From these crops, products such as cottonseed cakes and the residues of the crops are 

available as feed for livestock. In some areas of this zone farmers grow soybean and leguminous forage crops. 

The humid zone occupies 19 percent of SSA mostly in central and west Africa. The zone receives more than 

1 500 mm of rainfall annually which can sustain plants for 270 to 365 plant growing days. The zone is found at 

low latitudes north and south of the equator. Soils in this zone include Ferralsols (Oxisols), Acrisols (Ultisols) and 

Luvisols (Alfisols), the last of which are commonly encountered at the forest-savannah boundary. Vegetation 

consists of rain forest and derived savannahs with natural vegetation dominated by tall, closed forest which 

may be evergreen or semi-deciduous and which is often floristically rich. The herbaceous vegetation often 

contains large amounts of the major nutrients. 


246 


The soils are strongly weathered and hence have high levels of iron and aluminium oxides and low levels 

of phosphorous. The organic matter content is therefore generally low and the soils are fragile and easily 

degraded when the vegetative cover is lost. This zone has limited potential for livestock development, 

particularly because of the threat of the trypanosomiasis-transmitting tsetse fly. 

9.2.5 | Highlands zone 

The highland zone represents 5 percent of the land area of SSA, most of which is in eastern Africa, and half 

in Ethiopia. This zone occupies areas above 1 500 m altitude that have a mean daily temperature of less than 

20 ºC. The main highland areas in sub-Saharan Africa (SSA) are in Ethiopia, Kenya, Uganda, Rwanda, Burundi, 

western Zaire, Tanzania, Angola and Lesotho. There are also many other areas above the 1500 m contour and 

some of these afford tsetse-free grazing (e.g. Fouta Djallon and Bamenda). The highland areas vary in climate, 

topography, soils and land use. 

Topography varies from gently rolling hills to deeply incised valleys and steep slopes. Soils are sometimes 

deep and fertile Vertisols and Nitosols, but shallow soils of inherently low fertility are widespread. In many 

mountain grassland areas, soils only have a very shallow surface horizon that is fertile. Undisturbed upland 

areas are normally stable, although some soils exhibit ‘slumping’ even where undisturbed. Cultivating the socalled 

‘duplex’ soils 5 

and soils that form a surface crust on slopes results in high run-off. Unless soil conservation 

measures are taken and soils are sufficiently covered with vegetation, overland flow removes large amounts 

of soil. The zone receives bimodal rainfall (>1000 mm annually) and there are two growing seasons. Livestock 

rearing is widespread: farmers grow fodder, and animal traction is of increasing importance. Population 

pressure is encouraging crop–livestock integration, for which the cool highlands have high potential 

9.3 | General soil threats in the region 

The various threats to soil health and ecosystem services in SSA include: (1) erosion by water or wind; (2) 

loss of soil organic matter; (3) soil nutrient depletion; (4) loss of soil biodiversity; (5) soil contamination; (6) 

soil acidification; (7) salinization and sodification; (8) waterlogging; and (9) compaction, crusting and sealing/ 

capping (Mabogunje, 1995; Oldeman, 1991; Meadows and Hoffman, 2002; World Bank, 1997; IFPRI, 1999). 

9.3.1 | Erosion by water and wind 

About 77 percent of SSA is affected by erosion, with the most serious erosion areas in the Republic of 

South Africa, Sierra Leone, Guinea, Ghana, Liberia, Kenya, Zaire, Central African Republic, Ethiopia, Senegal, 

Mauritania, Nigeria, Niger, Sudan and Somalia. 

According to the GLASOD results (ISRIC/UNEP, 1990), about 494 million ha of the land in SSA is affected by 

one form of degradation or another. Of this, 227 million ha (46 percent) is by water erosion, 187 million ha (38 

percent) by wind erosion, 62 million ha (12 percent) by chemical degradation and 18 million ha (4 percent) by 

physical degradation. The intensity of water erosion has been described as very high to extreme on about 102 

million ha (45 percent of the total SSA area affected), moderate on about 67 million ha (30 percent) and slight 

on about 58 million ha (25 percent) (Oldeman, 1991). 

Water erosion: This is the most widespread soil degradation type in SSA. Water erosion increases on 

slopes where vegetation cover is reduced due to deforestation, overgrazing or cultivation that leaves the soil 

surface bare. It is further aggravated where there has been a loss of soil structure or infiltration rates have 

been reduced. The areas particularly affected are humid and sub-humid zones. Almost 70 percent of Uganda 

was degraded by soil erosion and soil nutrient depletion between 1945 and 1990. More than 20 percent of 

agricultural land and pastures in the country have been irreversibly degraded. 


247 


Water erosion poses the greatest threat to soils in Nigeria, where it affects over 80 percent of the land 

(NEST, 1991). Wind, sheet, gully and beach erosion affect different parts of the country at varying intensities, 

but attention will focus here on the impact of erosion on agricultural land. While wind erosion is confined 

to the arid north, sheet erosion by water is ubiquitous throughout the country. Areas most prone to sheet 

erosion are where farming has cleared the original vegetation, and the soils became impoverished scrubland. 

Gully erosion is by far the most alarming type of erosion, particularly in the Eastern region, because it often 

threatens settlements and roads. Although it affects a small fraction (less than 0.1 percent) of Nigeria’s 

924 000 km 2 of landmass, gully erosion claims large amounts of public funds annually for remedial action. 

Wind erosion: Wind erosion occurs most frequently in the arid and semi-arid parts of SSA, especially in 

areas with sandy or loamy soils. Wind erosion leads to loss of topsoil over extended areas causing soil fertility 

decline. Bielders, Michels and Rajot (1985) stated that wind erosion can remove up to 80 tonnes of soil from 

1 ha in a givenyear. In SSA wind erosion is second in importance to water erosion, constituting 38 percent of 

the total erosion in the region (ISRIC/UNEP, 1990) and affecting about 186 million ha of land in the region. The 

intensity of wind erosion is strong on about 9 million ha (5 percent), moderate on 89 million ha (48 percent) 

and light on 89 million ha (48 percent) (Oldeman, 1991). Over 99 percent of wind erosion in Africa occurs in the 

dry land zone, with less than 1 percent affecting the humid zone. 

Wind erosion is a natural process that commonly occurs in deserts and on coastal sand dunes and beaches. 

During drought, it can also occur in agricultural regions where vegetation cover is reduced. If the climate 

becomes drier or windier, wind erosion is likely to increase. Climate change forecasts suggest that wind 

erosion will increase over the next 30 years due to more droughts and more variable climate. The combination 

of a changing climate and consequent increase in wind erosion will cause a series of changes affecting soils: 

• less rain, which will support less vegetation 

• lower soil moisture, which will decrease the ability of soil particles to bind together into larger, heavier 

aggregates 

• increased wind speeds, which will result in more force exerted on the ground surface and more wind 

erosion (if wind speed doubles, the erosion rate increases eight times) 

• large losses of soil and nutrients 

• more large dust storms, which will impact soils and the community 

• poorer air quality, increased respiratory health risks, and temperature and rainfall changes due to 

atmospheric pollution (all off-site effects). 

9.3.2 | Loss of soil organic matter 

Land degradation leads to a release of carbon to the atmosphere through oxidation of soil organic matter 

(Oldeman, Hakkeling and Sombroek, 1991). Africa’s current major negative role in the global carbon cycle 

can be attributed to the substantial releases of carbon associated with land use conversion from forest or 

woodlands to agriculture (Smith, 2008). In the 1990s, these releases accounted for approximately 15 percent 

of the global net flux of carbon from land use changes (Hooper et al., 2006). Land management following 

conversion also impacts carbon status, soil fertility, and agricultural sustainability – a point underlined by 

many including Lal (2006), Ringius (2002), Zivin and Lipper (2008) and Tieszen, Tappan and Toure (2004). Soils 

often continue to lose carbon over time following land conversion (Woomer, Toure and Sall, 2004; Tschakert, 

Khouma and Sene, 2004; Liu et al., 2014), resulting in further reductions in crop yields and impoverishment 

of the farming population. However, these carbon stocks can be replenished with combinations of residue 

retention, manuring, nitrogen (N) fertilization, agroforestry, and conservation practices (Lal, 2006). 

In most sub-humid and semi-arid areas, much of the grazing land is burned annually during the dry season 

to remove the old and coarse vegetation and to encourage the growth of young and more nutritious grasses. 

Burning causes the loss of soil organic matter (released as CO 2 

) and thus impairs agricultural productivity. 


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It exposes the soil to the erosive forces of the wind during the dry season and of the rain during the rainy 

season. Furthermore, the annual burn of the vegetation severely reduces the return of organic matter to the 

soil. This results in loss of the benefits of soil organic matter, including fertility, structure, water retention and 

biodiversity. The soil becomes biologically, chemically and physically poorer (FAO, 2001). Land degradation 

further leads to a release of carbon to the atmosphere through the oxidation of soil organic matter which 

results from soil disturbance and from the consequent exposure of new soil surfaces to the weather. 

In agricultural land, the challenge has been to produce increasing quantities of food in an economic and 

institutional context where the means to improve productivity in a sustainable fashion are generally not 

available (e.g. lack of sustainable technological packages, absence of extension, training or affordable inputs 

etc.). Pressures to increase output in the absence of these supporting factors has led to: (i) the rapid expansion 

of agricultural land (over 65 percent in the last three decades); and (ii) the shortening of the fallow periods 

in traditional, extensive land use systems, which reduced the rehabilitation of soil fertility through natural 

processes. The increased use of fire as a clearing tool has led to the further loss of nutrients in many systems. 

Fertilizer consumption has not increased to compensate for the loss of soil nutrients resulting from the 

intensification of land use. Hence, there has been widespread mining of soil organic matter and nutrients. 

As a consequence of this poor land management combined with the vulnerable nature of many soils, much 

of SSA’s cropland is now characterized by low organic matter content, often in combination with a low pH 

and with aluminium toxicity. On degraded soils with low organic matter, inorganic fertilizers are also easily 

leached, which is likely to have negative long-term effects on agricultural productivity and on the quality of 

downstream water resources. 

9.3.3 | Soil nutrient depletion 

Soils in a large part of SSA are strongly weathered and inherently low in organic matter. Because of the 

increasing pressure on land, natural replenishment of nutrients during fallow periods is now insufficient 

to maintain soil productivity over the long-term. Insufficient nutrient replacement in agricultural systems 

on land with poor to moderate potential results in soil degradation. Already soil moisture stress inherently 

constrains land productivity on 85 percent of soils in Africa (Eswaran, Reich and Beinroth, 1997). Now soil 

fertility degradation places an additional serious human-induced limitation on productivity. 

The low nutrient status of most soils in SSA is further exacerbated by insufficient use of fertilizers and 

manure and by the practice of mono-cropping. Overall use of inorganic fertilizers in SSA is just 12 kg ha -1 , the 

lowest in the World, and soil nutrient depletion is widespread in croplands. Approximately 25 percent of soils 

in Africa are acidic, and therefore deficient in phosphorus (P), calcium and magnesium with often toxic levels 

of aluminium (McCann, 2005). Use of fertilizer in the region involves average applications of less than 9 kg of 

nitrogen and 6 kg of phosphorus per ha, compared with typical crop requirements of 60 kg of nitrogen and 30 

kg of phosphorus per ha. Recent research estimates that on average every country in SSA has a negative soil 

nutrient balance; in all countries studied, the amount of nitrogen, phosphorus and potassium (K) added as 

inputs was significantly less than the amount removed as harvest or lost by erosion and leaching (Swift and 

Shepherd, 2007). Although many farmers have developed soil management strategies to cope with the poor 

quality of their soil, low inputs of nutrients, including of organic matter, contribute to poor crop growth and 

to the depletion of soil nutrients. 

Stoorvogel, Smaling and Janssen (1993) calculated nutrient balances for arable soils in 38 sub-Saharan 

countries and for 35 crops for 1982 -1 983 and made forecasts for 2000. Subtracting values of the output (made 

up of harvest, removal of residues, leaching, denitrification and erosion) from the values of the input (made 

up of fertilizers, manures, rain, dust, biological N-fixation and sedimentation), the study reported alarming 

average nutrient losses for SSA as follows: 1982 -1 983: 22 kg N, 2.5 kg P and 15 kg K; 2000: 26 kg N, 3 kg P and 19 

kg K. This indicated persistent nutrient mining over time (Bationo et al., 2012). Other estimations claim that 

each year 4 million tonnes of nutrients are harvested annually in SSA against

Sub-national studies of nutrient depletion found annual losses of 112 kg per ha of N, 2.5 kg of P, and 70 kg 

of K in the western Kisii highlands of Kenya. Significantly lower losses were, however, recorded in southern 

Mali (Smaling, 1993; Smaling, Nandwa and Janssen, 1997). Farm monitoring and modelling of nutrient cycles 

for the western highlands of Kenya found that more nitrogen (63 kg per ha) was being lost through leaching, 

nitrification, and volatilization than through removal of crop harvests (43 kg per ha). Depending on the type 

of farm management practice, net nitrogen balances on cropped land varied between -39 and 110 kg per ha 

peryear, and net phosphorus balances between -7 and 31 kg per ha per year (Shepherd and Soule, 1998). 

9.3.4 | Loss of soil biodiversity 

Loss of soil biodiversity is considered the fourth major threat in SSA. Biodiversity loss occurs in a number 

of ways including destruction of habitat, land use change, introduction of new species, and harvesting and 

hunting of individual wild species. It has been estimated that in the mid -1 980s in Sub-Saharan Africa (SSA), 65 

percent of the ‘original’ ecosystems had been converted (Perrings and Lovett, 1999). The most important factors 

affecting soil biodiversity are: (i) habitat fragmentation; (ii) resource availability – the amount and quality of 

nutrients and energy sources; (iii) temporal heterogeneity e.g. seasonal effects; (iv) spatial heterogeneity - 

spatial differences in the soil; (v) climate variability; and (vi) interactions within the biotic community. 

Habitat destruction and/or fragmentation remains the primary threat to soil biodiversity in SSA. For 

instance, the once great equatorial forest that stretched from western Africa into eastern Africa is now 

fragmented into pockets represented by Lamto forest in Ivory Coast, Mbalmayo forest in Cameroun, Congo 

forest in Democratic Republic Congo, Kabale, Budongo and Mabira forests in Uganda and Kakamega forest 

in Kenya. The surrounding communities still rely heavily on these forests for basic needs such as fuelwood, 

charcoal, timber, poles, and other building materials. Due to human encroachments, the forests are subject 

to a mosaic of different land uses. There are patches of secondary forest, fallow and arable fields amidst 

significant remnants of primary vegetation. In the process of conversion and change in land use, soil biota 

have not been spared. Studies by Okwakol (2000) and Ayuke et al. (2011) have shown that up to 50 percent of 

soil macrofauna species within the forest area have been lost due to habitat destruction or fragmentation. 

Other threats to soil biodiversity in SSA include land use and land cover change, mainly through conversion 

of natural ecosystems, particularly forests and grasslands, to agricultural land and urban areas. In a study 

conducted across different ecosystems of Eastern (Kenya), Western (Nigeria, Burkina Faso, Ghana, Niger) and 

Southern Africa (Malawi), Ayuke et al. (2011) demonstrated a substantial reduction in the number of species 

and abundance of soil macrofauna groups such as earthworms and termites because of conversion of native 

or undisturbed ecosystems into arable systems. Continuous cultivation also exacerbates soil biodiversity loss 

because of loss of soil organic matter and hence of food resources for the soil organisms (Ayuke et al., 2011). It is 

likely that land clearing and deforestation will continue, further threatening genetic diversity as more species 

are lost (IAASTD, 2009). Sub-Saharan Africa suffers the world’s highest annual deforestation rate because of 

overexploitation of forest resources and conversion of forested land to agriculture. Although deforestation 

occurs throughout the continent, particularly affected areas are the moist forests of Western Africa and the 

highland forests of the Horn of Africa (FAO, 2007; Hansen et al., 2013). 

Mulugeta (2004) reported that in Ethiopia deforestation and subsequent cultivation of the tropical dry 

Afromontane forest soils endangered the native forest biodiversity not only through the outright loss of 

habitat but also by impairing the soil seed banks. The results showed that the contribution of woody species 

to the soil seed bank declined from 5.7 percent after sevenyears to nil after 53 years of continuous cultivation. 

However, soil quality and native flora degradation are reversible through reforestation. In fact, reforestation 

of abandoned farm fields with fast-growing tree species was shown to restore soil quality. Tree plantations 

established on degraded sites also fostered the recolonization of diverse native forest flora under their 

canopies. An important result from studying the effects of reforestation is that good silviculture, particularly 

selection of appropriate tree species, can significantly affect the rate and magnitude of restoration processes 

for both soil quality and biodiversity. 


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In many African cultures, harvesting of soil fauna groups such as the termite alates and queens, chafer 

grubs for food, and the use of earthworms as bait by fishermen can also be a threat to soil biodiversity, and 

may in the long run contribute to substantial loss of many species of soil fauna. 

Harsh climatic conditions and/or climate change may also contribute to changes in soil biodiversity in SSA. 

For example, a more than average numbers of earthworm and termite taxa are found under relatively warmer, 

drier conditions (Ayuke et al., 2011). This is contrary to the observation that earthworm and termite diversity 

increases with increases in rainfall or soil moisture, as generally found in temperate climates (Bohlen et al., 

1995; Curry, 2004). However, seasonality of rainfall in the tropical regions means rainfall amounts per season 

may be more important than the annual total. Lower taxonomic richness among sites in Eastern Africa may be 

attributable to less favourable conditions arising from high rainfall and low temperatures at higher altitudes 

(Ayuke et al., 2011). 

Intense management practices that include application of pesticides and frequent cultivation affect soil 

organisms, often altering community composition of soil fauna. Soil biological and physical properties (e.g., 

temperature, pH, and water-holding characteristics) and microhabitat are altered when natural habitat is 

converted for agricultural production (Crossley, Mueller and Perdue, 1992). Changes in these soil properties 

may be reflected in the distribution and diversity of soil meso fauna. Organisms adapted to high levels of 

physical disturbance become dominant within agricultural communities, thereby reducing the richness and 

diversity of soil fauna (Paoletti, Foissner and Coleman, 1993). 

The extent of soil sterilization and loss of soil biodiversity in SSA has yet to be quantified on a large-scale 

across the region. However, it is clear that unsustainable soil management practices have depleted soil 

organic matter, promoted soil degradation and may have caused soil fauna and flora imbalances. This land 

degradation will continue unless land users in SSA adopt an agro-biological approach to managing their soils 

(Van der Merwe et al., 2002). 

9.3.5 | Soil contamination and pollution 

Chemical fertilizers and pesticides have had negative effects on the environment in most SSA countries. 

However, soil pollution through agrochemical use in SSA has been of less concern compared to other regions of 

the world, mainly because of the low levels of application. However, with the increasing push towards higher 

use of fertilizer, pesticide and herbicide to boost productivity, efforts will be needed to reduce the associated 

negative impacts on soil quality (IAASTD, 2009). 

Chemical pollution has emerged as a threat to soil quality. According to a United Nations Environment 

Programme (UNEP, 2007) environmental assessment in ten communities in Ogoni land in southeastern 

Nigeria which had been affected by crude oil spills, drinking water, the air and agricultural soils contained 

over 900 times the permissible levels of hydrocarbon and heavy metals. The report acknowledged that, even 

if all its recommendations were implemented, recovery might take 30 years. Other published research work 

suggests that heavy metal pollution is occurring across SSA. Heavy metal (Pb, Cd, Hg, Cu, Co, Zn, Cr, Ni, As) 

pollution of soils has been reported in Nigeria, Kenya, Ghana and Angola (Fakayode and Onianwa, 2002; 

UNEP, 2007; Odai et al, 2008). 

Change of land use, particularly urbanization, is another factor in soil contamination. National data from 

South Africa indicate that areas under urban, forestry and mining land uses have all increased over the last 

decade, whereas the cultivated area has decreased. The urban area has increased from 0.8 percent of total 

area to 2 percent, forestry from 1.2 percent to 1.6 percent, and mining from 0.1 percent to 0.2 percent, while 

the cultivated area has decreased from 12.4 percent to 11.9 percent. The increase in the urban and mining areas 

is a major concern in terms of soil conservation and future use. Urban development involves soil sealing which 

irreversibly removes soils from other land uses. Mining results in serious chemical and physical soil degradation 

which subsequently can only be partially restored. 


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9.3.6 | Soil acidification 

In SSA, extremely acid soils, mainly potential or actual acid sulphate soils, occur only in a small area around 

the Niger delta and sporadically along the coastal plains of West Africa. Other acid soils occupy about 15 

percent of the continent and are mainly found in the moist parts of the semi-arid zones and in sub-humid 

areas. Many of the Acrisols (Ultisols) and some Lixisols (Alfisols) have acid surface and subsurface horizons 

which, coupled with the moisture stress conditions, makes these soils extremely difficult to manage under 

low-input conditions. In West Africa, the annual additions of dust from the Sahara brought by the Harmattan 

winds raise the pH of the surface horizons. The problem is therefore less acute there, although subsoil acidity 

remains (Eswaran et al., 1996). Another region of acid soils occurs south of the tropic of Capricorn and includes 

parts of South Africa (Beukes, Stronkhorst and Jezile, 2008a,b) where it poses a serious soil chemical problem 

and is in fact one of the greatest production-limiting factors. 

9.3.7 | Salinization and sodification 

Salinization is defined as a change in the salinity status of the soil. This can be caused by improper 

management of irrigation schemes, particularly in the arid and semi-arid regions. Irrigation-induced soil acidity 

is aggravated when irrigation is practiced on soils unsuitable for irrigation (Barnard et al., 2002). Salinization 

can also be caused if sea water intrudes into coastal regions either on the surface or into groundwater. It may 

also arise in closed basins when there is excessive abstraction of groundwater from aquifers of different salt 

content. Salinization also takes place where human activities lead to increased evapotranspiration from soils 

on salt-containing parent material or where saline ground water is being pumped out (Oldeman, 2002). 

In the arid and semi-arid parts of Africa, soil salinity and alkalinity are major problems affecting about 24 

percent of the continent. Soils with pH>8.5 are designated as alkaline (Eswaran et al., 1996). Soil salinity and 

sodicity problems are common in arid and semi-arid regions where rainfall is insufficient to leach salts and 

excess sodium ions out of the rhizosphere. More than 80 million ha of such soils are found in Africa. 

Increasing temperatures may result in high evaporative demands that may activate the capillary rise of 

salts, leading to soil salinization. The results of a study in Sudan showed a significant increase in salinity in 

the Dongla area in the north, where the annual rainfall is the lowest in the country. This increase is associated 

with fluctuation and erratic distribution of rainfall, as well as with a rise in temperature (Abdalla et al., 2011). 

9.3.8 | Waterlogging 

Human intervention in natural drainage systems may lead to waterlogging or flooding by river water. Most 

waterlogging threats are due to effects of human-induced hydromorphy. Causes include a rising water table 

(for example, due to construction of reservoirs or irrigation) or increased flooding caused by higher peak flows 

of rivers. The technology of flooding in paddy fields to provide a proper environment for paddy rice is generally 

not considered a threat to ecosystem services, although it may increase the emissions of GHG. It is estimated 

(Oldeman, Hakkeling and Sombroek, 1991) that waterlogging constitutes 1.5 percent of the non-erosion soil 

degradation threats in Africa. 

9.3.9 | Compaction, crusting and sealing 

The population of the Sub-Saharan Africa (830 millions) is approximately 12 percent of the world population. 

SSA population has been growing at a rate of 2.6 percent year during the last decade, although the rate is now 

declining. The tendency in the region is towards the concentration of growing populations in moderately large 

cities (rather than mega-cities). Since the early 1970s, several SSA countries have experienced accelerated urban 

expansion, recording some of the highest urban growth rates in the world of up to 5 percent per year (Todaro, 

2000). There are numerous examples of single-city dominance in the region. For instance, in Mozambique, 


252 


Maputo accounts for 83 percent of the country’s urban population, while the figures for Dakar, Lome, Kampala 

and Harare are 65, 60, 52 and 50 percent respectively (World Bank, 2002). Nigeria and South Africa represent 

exceptions to this single-city dominance, as they have several large and well distributed urban centres. South 

Africa and Nigeria are also the countries recording the highest amount of impervious surface area (ISA) in 

the region, and they have high Urbanization Indexes (the ratio between the total area of the country and the 

urbanized area) (Figure 9.2). 

9.4 | The most important soil threats in Sub-Saharan Africa 

Of the threats to soils and related ecosystem functions in SSA listed in Section 9.3, the most critical are soil 

erosion, loss of soil organic matter and soil nutrient depletion (UNEP, 2013). Loss of soil biodiversity is also a 

significant threat in SSA. These four threats are interrelated. More is known about the first three and these are 

discussed in greater detail in this section. 

10 

8 

Urban Area (SQKM) 

Urbanization index (%) 

1,2 

1 

Urban Area (thousands SQKM) 

6 

4 

0,8 

0,6 

0,4 

2 

0,2 

0 

0 

South Africa 

Nigeria 

Sudan 

Democratic Republic of the Congo 

Morocco 

Cote d'Ivoire 

Ghana 

Mozambique 

United Republic of Tanzania 

Ethiopia 

Cameroon 

Zambia 

Angola 

Niger 

Kenya 

Senegal 

Zimbabwe 

Uganda 

Mali 

Republic of the Congo 

Benin 

Botswana 

Burkina Faso 

Namibia 

Gabon 

Madagascar 

Chad 

Central African Republic 

Togo 

Somalia 

Sierra Leone 

Mauritania 

Liberia 

Rwanda 

Malawi 

The Gambia 

Eritrea 

Lesotho 

Burundi 

Guinea-Bissau 

Swaziland 

Equatorial Guinea 

Mauritius 

Djibouti 

Reunion 


Figure 9.2 Extent of urban areas and Urbanization Indexes for the Sub-Saharan African countries. Source: Schneider, Friedl and 

Potere, 2010. 


253 



Direct causes of soil erosion 

Expansion of land for agriculture: Soil erosion can be a natural process but it is also often caused or 

accelerated by human activities that involve inappropriate land use. As discussed above (Section 9.3), much 

of the change in land use practices in SSA has been driven by the need to increase production and incomes 

in an economic and institutional context where the means to improve productivity in a sustainable fashion 

are generally not available. The processes involved - rapid expansion of agricultural land, shortening of fallow 

periods, increased use of fire – are discussed in full in Section 9.3.2 above. 

As fertilizer consumption did not increase to compensate for the loss of soil nutrients resulting from the 

intensification of land use, there has been widespread mining of soil nutrients and soil organic matter. As a 

consequence of the type of soils that occur in the region and because of generally poor land management, 

many SSA croplands now have low soil organic matter contents and soils that have a low pH and suffer from 

aluminium toxicity. On degraded soils with low organic matter, inorganic fertilizers are also easily leached, and 

this process has devastating long-term effects for agricultural productivity. Alternative means of maintaining 

soil fertility, such as crop rotation with biological nitrogen fixing (BNF) species, application of green manure, 

agroforestry, composting, rock phosphates, etc., have proved to be highly effective at the local scale. However, 

these technologies have not been applied widely enough to have an impact at a national let alone continental 

scale. 

Overgrazing: There has been much debate on the impacts in SSA of high grazing pressures on vegetation 

composition in rangelands. The current understanding is that continued high grazing pressure may affect 

rangeland productivity, particularly in the long term. Vegetation studies also show that high grazing pressures 

lead to changes in species composition, which may reduce the resilience of rangelands to drought (Hein and 

De Ridder, 2006). During a drought, degraded rangelands show a much stronger decline in productivity 

than non-degraded rangelands. Recentyears have seen droughts with severe impacts on livestock and local 

livelihoods in parts of Niger and in the East African drylands (Uganda and Kenya). 

Deforestation: Most forests and woodlands in SSA are experiencing rapid rates of deforestation. 

Deforestation is driven by a number of processes, in particular: (i) the continued demand for agricultural land; 

(ii) local use of wood for fuel, charcoal production and construction purposes; (iii) large-scale timber logging, 

often without effective institutional control of harvest rates and logging methods; and (iv) population 

movements and resettlement schemes in forested areas. The amount of cropped land in SSA has increased by 

about 40 million ha in 30 years (1975-2005), most of it at the expense of forests and woodlands (FAO, 2015). 

Further expansion of cropland would be at the expense of forests or rangeland. 

Socio-economic causes of soil erosion 

Population expansion: Behind these direct drivers of erosion lies the demographic driver of a continuously 

growing population (Figure 9.3). The rate of SSA population growth has moderated in recentyears and is 

currently 2.1 percent per year. Nonetheless, in the next 15 years SSA will have to accommodate at least 250 

million additional people, a 33 percent increase (UNDP, 2005). With the increase in population comes an 

increased demand for living space and food which will directly affect soil use in the region. 


254 


Poverty: General poverty of the farming population and the low potential of the farming systems 

characteristic of SSA pose considerable challenges to sustainable agricultural growth and poverty reduction. 

Poverty is particularly prevalent in the pastoral/agro-pastoral, highland perennial and forest based farming 

systems which constitute one-third of the total SSA production systems (FAO and World Bank, 2001). From 

Figure 9.4 it is clear that SSA has many countries where a large percentage of the population is living below 

the poverty line. 

Figure 9.3: The fertility rate (the number of children a woman is expected to bear during her lifetime) for 1970 and 2005. Source: 

Fooddesert.org 

Climate Change: Climate change is predicted to affect SSA agro-ecosystems on a significant scale in the 

coming decades. The continent has a long history of rainfall fluctuations of varying lengths and intensities. 

Severe droughts affected East and West Africa alike during the 1910s. Drought episodes were generally followed 

by increasing rainfall levels, but negative trends were observed again from the 1950 onwards, culminating in 

the droughts of the early 1970s and mid -1 980s. These droughts had an impact on the susceptibility of soils to 

erosion. 

Legend 

Population (%) 

3 - 14 

14 - 26 

26 - 37 

37 - 48 

48 - 57 

57 - 66 

66 - 75 

75 - 84 

Figure 9.4 Percentage of population living below the poverty line. Source: CIA World Factbook, 2012. 


255 


Water erosion: its extent and distribution in the region 

Water erosion constitutes 46 percent of the land degradation types in SSA, and wind erosion accounts for 

a further 38 percent (FAO, 2005). The most recent continent-wide assessment shows that 494 million ha, or 

22 percent of the agricultural land (including rangelands) in Africa, are affected by water erosion (Oldeman, 

Hakkeling and Sombroek, 1991). The assessment confirms common field observations that overgrazing is the 

main cause of soil erosion, followed by inappropriate cultivation techniques on arable land. In this context 

it is important to note that the number of cattle in Africa almost doubled in the period 1961 -1 994, while the 

area of grazing lands hardly increased (FAO, 2015). For the future, the expected intensification of use on 

currently cultivated lands, expansion of cultivation into more marginal areas, reduction in grazing lands and 

the increasing numbers of livestock are likely to increase vulnerability to erosion. 

As discussed above (9.3.1), severely eroded areas in Africa can be found in South Africa, Sierra Leone, Guinea, 

Ghana, Liberia, Kenya, Nigeria, Zaire, Central African Republic, Ethiopia, Senegal, Mauritania, Niger, Sudan 

and Somalia. More than 20 percent of SSA’s agricultural land and pasture has been irreversibly degraded, 

mainly by soil erosion (UNSO/SEED/BDP, 1999). 

Erosion has assumed a serious dimension in Nigeria, affecting every part of the country. In the eastern part 

of the country, erosion has ravaged wide areas. Active and inactive gullies have formed with surface areas 

ranging from 0.7 km 2 in Ohafia to 1.15 km 2 in Abiriba in Abia State. The width of the gullies ranges between 0.4 

km in Ohafia and 2.4 km in Abiriba. A gully with a depth of 120 m has been recorded at Abiriba (Ofomata, 1985). 

In addition, agricultural practices have contributed to the problems of widespread sheet erosion. Erosion is 

thus exerting major pressure on soil resources with far-reaching consequences for both the population and 

the environment (Jimoh, 2000). In the northern areas of Nigeria, erosion is equally serious, especially in 

places like Shendam and Western Pankshin in Plateau State, as well as at Ankpa and Okene in Kogi State. 

Gully erosion is also prominent in Efon-Alaaye, Ekiti State in the western part of the country (Adeniran, 1993). 

The areas of Nigeria most affected by erosion are the Agulu and Nanka districts of the eastern part of 

Nigeria, and the Shendam and western Pankshin areas of Plateau State, Nigeria (Udo, 1970; Okigbo, 1977). 

Elsewhere, the Imo State government has estimated that about 120 000 km 2 of land has been devastated 

by gully erosion. As a result, eight villages have been destroyed and 30 000 people needed to be resettled. 

Erosion damage in Imo and Anambra states was estimated to cause the loss of over 20 tonnes of fertile soil per 

annum, at an economic cost of over 300 million naira per annum. Gullies extended to depths of over 120 m and 

widths up to 2 km wide (Adeleke and Leong, 1980). In 1994, about 5 000 people were rendered homeless due 

to erosion in Katsina State, Nigeria. Properties worth over 400 million naira were destroyed and many lives 

lost. Other areas affected by erosion include Auchi in Edo State, Efon Alaye in Ondo State, Ankpa and Okene in 

Kogi State, and Gombe in Bauchi State. In many areas, erosion has resulted in a physical loss of available land 

for cultivation. For example, about 1 000 ha of cultivable land has been lost to erosion at the Agulu-Nanka 

area of Nigeria. Thus the loss of homes and crops, disruption of communication routes, financial losses and 

attendant hydrological problems can all stem from erosion problems. 

Nearly 90 percent of rangelands and 80 percent of farmlands in the West African Sahel, Sudan, and 

northeast Ethiopia are seriously affected by land degradation, including soil erosion. More than 25 percent 

of South Africa is seriously degraded by erosion. Almost 70 percent of Uganda’s territory was degraded by soil 

erosion and soil nutrient depletion between 1945 and 1990. Across SSA, more than 20 percent of agricultural 

land and pastures has been irreversibly degraded, affecting more than 65 percent of Africa’s population (Global 

HarvestChoice, 2011). 


256 


Considering that over 80 percent of South Africa’s land surface is covered by natural vegetation, the 

estimated annual soil loss of 2.5 tonnes soil per ha is excessive. These rates of soil loss far exceed tolerance 

levels and are almost ten times the estimated rate of soil formation, which has been estimated at 0.31 tonnes 

ha -1 yr -1 in the case of a 1 m thick solum of a tropical soil (Van der Merwe, 1995; see Section 6.1). Soil organic 

matter (SOM) plays a major role in ensuring soil stability. There is a general decline in SOM in South African 

soils. An estimated 20 percent of the country’s total surface area is potentially highly erodible. Bearing in mind 

the country’s geology, rainfall and topographic characteristics in addition to declining SOM, soil erosion is 

likely to stay a dominant soil degradation process. 

Sediment movement by erosion contributes significantly to shifts in soil fertility. Sediment movement is 

widespread in South Africa as reflected by the annual losses of 3 300 tonnes N, 26 400 tonnes P and 363 000 

tonnes K estimated by Du Plessis in 1986 (Van der Merwe, 1995). Periodic floods transport massive amounts of 

sediment and nutrients within catchments. The Demoina flood in 1984, for instance, deposited as much as 34 

million tonnes of sediment in the Mfolozi flats (Scotney and Dijkhuis, 1990). One 1985 study used a siltation 

approach to estimate the siltation load carried by the Tugela River, finding soil loss from the catchment area 

as high as 4.4 tonnes ha -1 yr -1 (De Villiers et al., 2002). It has been estimated that water erosion affects 6.1 

million ha of cultivated soils in South Africa. Of this area, 15 percent of soils are seriously affected, 37 percent 

moderately affected, and the rest slightly affected. 

Wind erosion in the region 

Wind erosion physically removes the lighter, less dense soil constituents such as organic matter, clays and 

silts, thus removing the most fertile part of the soil and lowering soil productivity (Lyles, 1975). In SSA, soil 

erosion by wind occurs mainly in the arid and semiarid regions. The occurrence of wind erosion at any one site 

is a function of weather events interacting with soil and land management through the effects of weather 

on soil structure, tilth and vegetation cover. At the southern fringe of the Sahara Desert, a special dry and 

hot wind, locally termed Harmattan, occurs. These North-easterly or Easterly winds normally blow in the dry 

winter season under a high atmospheric pressure system. When the wind force of Harmattan is beyond the 

threshold value, sand particles and dust particles will be blown away from the land surface and transported 

for several hundreds of kilometres across the land and as far as the Atlantic Ocean (WMO, 2005). Areas in SSA 

most susceptible to wind erosion are the southern fringe areas of the Sahara, Botswana, Namibia, Zimbabwe, 

Tanzania and South Africa (Favis-Mortlock, 2005). 

It is estimated that 25 percent of South Africa is affected by wind erosion (Laker, 2005), amounting to an 

estimated 10.9 million ha. Of this area, 7 percent is seriously affected, 29 percent moderately and 64 percent 

slightly (Barnard et al., 2002). Wind erosion is particularly evident on drift sands in the coastal areas, but also 

on cultivated land in the Highveld areas. The seriousness of wind erosion can be deduced from the situation in 

the Eastern Cape Province where there are over 14 000 ha of drift sand (Barnard et al., 2002). Most of South 

Africa’s prime agricultural soils in the relatively arid western part of the country are wind-blown sand deposits 

(De Villiers et al., 2002). 

Wind erosion may cause off-site effects, such as the covering of the terrain with wind-borne soil particles 

from distant sources. It is estimated that more than 100 million tonnes of dust per annum are blown westward 

from the African continent across the Atlantic. The amount of dust arising from the Sahel zone has been 

reported to be around or above 270 million tonnes per year which corresponds to a loss of 30 mm per m 2 

per 

year or a layer of 20 mm of soil particles over the entire area (WMO, 2005). 


257 


9.4.2 | Loss of soil organic matter 

The loss of vegetative cover and decline in the level of soil organic matter (SOM) are the root cause of 

most soil degradation, since all the physical, chemical and biological problems follow a drop in SOM content. 

Soil organic matter is a key component of any terrestrial ecosystem, and any variation in its abundance and 

composition has important effects on many of the processes that occur within the system. The amount 

of organic matter and size of soil carbon stock results from an equilibrium between the inputs into the 

system, which are mostly from biomass detritus, and outputs from the system, largely decomposition and 

volatilization. These processes are driven by various parameters of natural or human origins (Schlesinger 

and Winkler, 2000; Amundson, 2001; Section 2.1). A decrease of organic matter in topsoil can have dramatic 

negative effects on the water holding capacity of the soil, on the soil structure stability and compactness, 

on nutrient storage and supply, and on soil biological components such as mycorrhizas and nitrogen-fixing 

bacteria (Sombroek, Nachtergaele and Hebel, 1993). 

Direct causes of SOM decline 

Apart from climatic factors that influence carbon changes in the soil, inappropriate land uses and practices 

are the main cause of decline in SOM. These uses and practices include: monoculture cereal production; 

intensive tillage; short to no fallow; and reduction or absence of crop rotation systems. The long-term effects 

of these management actions are now being experienced across SSA. 

The SSA experience is not unique. Across the globe, the carbon balance of terrestrial ecosystems is being 

changed markedly by the direct impact of human activities. Land use change was responsible for 20 percent 

of global anthropogenic CO 2 

emissions during the 1990s (IPCC, 2007). In SSA, land use change is the primary 

source, much of it through burning of forests. 

The impact of land use change varies according to the land use types. The clearing of forests or woodlands 

and their conversion into farmland in tropical SSA reduces the soil carbon content mainly through reduced 

production of organic inputs, increased erosion rates and the accelerated decomposition of soil organic matter 

by oxidation. Various reviews agree that the loss amounts to 20 to 50 percent of the original carbon in the 

topsoil, with deeper layers less affected, if at all (Sombroek, Nachtergaele and Hebel, 1993; Murty et al., 2002; 

Guo and Gifford, 2002). However, conversion of forests to pasture does not necessarily change soil carbon 

(Guo and Gifford, 2002) and may actually increase the soil organic matter content (Sombroek, Nachtergaele 

and Hebel, 1993). Where shifting cultivation is practiced, soil carbon has been found to reduce to half the level 

before the land was cleared for use (Detwiler, 1986). Surprisingly, studies suggest that commercial logging and 

tree harvesting do not result in long-term decreases in soil organic matter (Knoepp and Swank, 1997; Houghton 

et al., 2001; Yanai et al., 2003). Clearly many factors are at play: changes in the amount of soil organic matter 

following conversion of natural forests to other land uses depend on several factors such as the type of forest 

ecosystem undergoing change (Rhoades, Eckert and Coleman, 2000), the post conversion land management 

practiced, the climate (Pastor and Post, 1986) and the soil type and texture (Schjønning et al., 1999). 

Socio-economic causes of SOM decline 

In Sub-Saharan Africa, socio-economic pressures to increase production and incomes create incentives for 

farmers to reduce the length of fallow periods, cultivate continuously, overgraze fields, or remove much of 

the above-ground biomass for fuel, animal fodder and building materials. These practices can result in the 

reduction of SOM, water holding capacity and nutrients. They also increase the soil’s vulnerability to erosion 

(Lal, 2004). 


258 


Extent of SOM decline in the region 

As with negative nutrient balances (see below, 9.4.3), SOM decline threatens soil productivity. In SSA, the 

concentration of organic carbon in the top soil is reported to average 12 mg kg -1 for the humid zone, 7 mg kg -1 

for the sub-humid zone and 4 mg kg -1 or less in the semi-arid zone (Williams et al., 1993). These inherently low 

concentrations of soil organic carbon are due not only to the low root growth of crops and natural vegetation 

but also to the rapid turnover rates of organic materials caused by high soil temperature associated with 

abundant micro-fauna, particularly termites (Bationo et al. 2003). There is considerable evidence for rapid 

decline in SSA of soil organic C levels where cultivation of crops is continuous (Bationo et al., 1995). For sandy 

soils, average annual losses in soil organic C may be as high as 5 percent, whereas for sandy loam soils reported 

losses are much lower, averaging 2 percent (Pieri, 1989). Results from long-term soil fertility trials indicate that 

losses of up to 0.69 tonnes carbon ha -1 yr -1 in the soil surface layers are common in Africa, even with high levels 

of organic inputs (Nandwa, 2003; Bationo et al., 2012). 

Responses to SOM decline 

Appropriate land management could reverse the trend of SOM decline and contribute to soil carbon 

sequestration. In fact, increasing the SOM content is crucial for future African agriculture and food production 

(Bationo et al., 2007; Sanchez, 2000). Several studies on SSA have shown that a synergetic effect exists 

between mineral fertilizers and organic amendments and that this synergy leads to both higher yields and 

higher SOC content (Palm et al., 2001, Vågen et al., 2005; Bationo et al., 2007). 

Barnard et al. (2002) emphasized the importance of establishing and maintaining an effective and intimate 

association between soils and growing plants. Biological measures for stabilizing slopes and decreasing the 

rate of runoff are essential. It is often necessary to undertake some form of land shaping prior to this, together 

with chemical amelioration and nutrient augmentation. 

There is abundant evidence that soil organic matter plays a major role in stabilizing soil and in preventing its 

physical, chemical and biological deterioration. This has been demonstrated under South African conditions 

by several scientists as reported by Barnard et al. (2002). For example, Folscher (1984) pointed out that microorganisms 

played a vital role in the chemo-biological condition of soils. Under predominantly heterotrophic 

microbial activity, physical and chemical stability could be expected, while under autotrophic microbial 

activity, acidification and nutrient decline could be forecast. Much more attention therefore needs to be paid 

to the dynamic nature of soil and its physical, chemical and biological interactions. 

Because nitrogen dynamics are so important in establishing a stable C:N ratio in soil, alternative sources of 

natural forms of nitrogen such as suitable legumes should be included in rotations. Rhizobial and mycorrhizal 

associations need to be stimulated and soil organic carbon and nutrient levels need to be systematically 

monitored and evaluated. Other soil quality indicators relating to specific situations need to be developed 

and utilized, with emphasis on earthworm populations as an indicator of soil quality. Reduced, minimum and 

no-till systems also need to be investigated and implemented where possible. These have been introduced 

worldwide and are being adopted in many parts of South Africa (Van der Merwe et al., 2000). 

Land degradation leads to a release of carbon to the atmosphere through oxidation of soil organic matter. 

With present concerns about climate change and the increase in atmospheric CO 2 

, it has been suggested that 

this process could be reversed and that the soil could be used to capture and store carbon. Soil organic matter 

could be gradually built up again through carbon sequestration. Among the land use changes which could be 

promoted with this objective in mind are improved agricultural practices, the introduction of agroforestry, and 

reclamation of degraded land. By such means, the carbon stored in soils could be substantially increased by 

amounts of the order of 30–50 tonnes ha -1 . Thus land use changes which are beneficial to local communities 

would, in addition, fulfil a global environmental objective. 


259 


9.4.3 | Soil nutrient depletion 

Nearly 3.3 percent of agricultural gross domestic product (Agricultural GDP) in Sub-Saharan Africa is lost 

annually due to soil and nutrient losses (Global HarvestChoice, 2011). In Africa, harvesting grains and crop 

residues from the land removes considerable quantities of soil carbon content. As lost nutrients in SSA are only 

very partially replaced with fertilizers, these losses contribute to negative nutrient balances (Gray, 2005). As 

a result, soil fertility decline has been described as the single most important constraint on food production 

and food security in SSA. Soil fertility decline (also described as soil productivity decline) is a deterioration 

of chemical, physical and biological soil properties. Besides soil erosion, the main processes contributing to 

nutrient depletion in SSA are: 

• Decline in organic matter and soil biological activity 

• Degradation of soil structure and loss of other soil physical qualities 

• Reduction in availability of major nutrients (N, P, K) and micro-nutrients 

• Increase in toxicity, due to acidification or pollution 

In the first assessment of the state of nutrient depletion in SSA, which was carried out in 1990, nutrient 

balances were calculated for the arable lands of 38 countries across the continent. Four classes of nutrient-loss 

rates were established (Table 9.2). As discussed above (9.3.3), the average nutrient loss in 1990 was estimated 

to be 24 kg nutrients ha -1 per year (10 kg N; 4 kg P 2 

O 5 

, 10 kg K 2 

O). Countries with the highest depletion rates, 

such as Kenya and Ethiopia (Table 9.3), also have severe soil erosion. 

Class Low Moderate High Very High 

N 40 

P 2 

O 5 

15 

K 2 

O 40 

Table 9.2 Classes of nutrient loss rate (kg ha -1 yr -1 ). Source: Stoorvogel and Smaling, 1990. 

Direct causes of nutrient decline 

Fertility decline is caused by a negative balance between output (harvesting, burning, leaching, and so 

on) and input of nutrients and organic matter (manure/fertilizers, returned crop residues, mineral deposition 

through flooding). The estimate of nutrient depletion in SSA cited above is worrying. However, some scientists 

(Roy et al., 2003) have expressed concern about the approach used, as it is based on approximation and 

aggregation at country level which could be misleading, masking the ‘bright’ spots and the ‘hot’ spots where 

urgent nutrient replenishment is required. Assessment of fertility decline at micro-watershed or community 

level would be more appropriate. 

Socio-economic causes of nutrient decline 

There are various factors that indirectly influence nutrient depletion in SSA and they vary between 

ecological regions and amongst countries and locations within a given ecological region. The cost of buying 

mineral fertilizer can put it beyond the reach of many SSA smallholders (World Bank, 1998). Farm-level 

fertilizer prices in Africa are among the highest in the world (Bationo et al. 2012). One metric tonne of urea, for 

example, costs about US$ 90 in Europe, US$ 500 in Western Kenya and US$ 700 in Malawi. These high prices 

can be attributed to the removal of subsidies, high transaction costs, poor infrastructure, and poor market 

development, inadequate access to foreign exchange and credit facilities, transportation costs and lack of 

training to promote and utilize fertilizers. For example, it costs about US$ 15, US$ 30 and US$ 100 to move 1 

tonne of fertilizer 1 000 km in the United States, India and SSA respectively (Bationo et al., 2012). 


260 


Many farmers do not follow recommended fertilizer application rates because of cash or labour constraints. 

In much of Southern and Western Africa there is a dry season lasting 7 to 8 months. In the first weeks of 

the rainy season many farm operations such as planting, weeding, and fertilizing must take place in rapid 

succession. Farmers who weed maize twice at critical periods can achieve a higher yield with half the amount 

of fertilizer used by farmers who only weed once (Kabambe and Kumwenda, 1995), but many farmers do not 

have sufficient labour to weed more often. 

Output price instability is an important factor posing risks for fertilizer users in Western Africa (Byerlee et 

al., 1994). When markets are sparse, as they are in many rural areas dominated by subsistence production, the 

variations in market prices of crops tend to be wider than in regions where markets are more fully developed. 

Overall, the economics of fertilizer use are often not sufficiently positive, especially under rainfed conditions; 

farmers are cash-poor and so cannot buy expensive inputs; and farmers are highly averse to making cash 

outlays in unpredictable climatic conditions and with uncertain commercial returns. 

Extent of nutrient decline in the region 

The results of an FAO study (1983-2000) (Lesschen et al., 2003; Stoorvogel and Smaling, 1990) which assessed 

N, P and K balances by land use system and by country revealed a generally downward trend in soil fertility 

in Africa. Overall, the study suggests that all African countries except Mauritius, Reunion and Libya show 

negative nutrient balances every year. The result for 2000 showed a deteriorating nutrient balance for almost 

all countries. This was influenced by the FAO estimates for crop production in 2000 and an accompanying 

expected decrease in fallow areas. For SSA as a whole, the nutrient balances were: -22 kg ha -1 in 1983 and -26 kg 

ha -1 in 2000 for N; -2.5 kg ha -1 in 1983 and -3.0 in 2000 for P; and -1 5 kg ha -1 in 1983 and -1 9 kg ha -1 in 2000 for K. 

Table 9.3 lists nutrient balances for several SSA countries. The study found substantial differences between 

countries. In 1993-95 the difference between nutrient inputs and nutrient losses in the continent ranged from 

-1 

4 kg of NPK per ha per year in South Africa to 136 kg in Rwanda. Burundi and Malawi also experienced rates of 

nutrient depletion above 100 kg of NPK per ha per year. 

Densely populated and hilly countries in the Rift Valley area (Kenya, Ethiopia, Rwanda and Malawi) had the 

most negative values, owing to high ratios of cultivated land to total arable land, relatively high crop yields, 

and significant erosion problems. In the semiarid, arid, and Sudano-Sahelian areas that are more densely 

populated, soils were found to lose 60 -1 00 kg of nitrogen, phosphorus, and potassium (NPK) per ha each year. 

The soils of these areas are shallow, highly weathered, and subject to intensive cultivation with low-level 

fertilizer use. 

Short growing seasons contribute to additional pressure on the land. In important agricultural areas in 

the sub-humid and humid regions and in the savannas and forest areas, nutrient losses vary greatly. Rates of 

nutrient depletion range from moderate (30 - 60 kg of NPK per ha per year) in the humid forests and wetlands 

in southern Central Africa, to high (> 60 kg NPK per ha per year) in the East African highlands. 

More countries fall into the high depletion range than the medium range. Nutrient imbalances are highest 

where fertilizer use is particularly low and nutrient loss, mainly from soil erosion, is high. The low gains in 

nutrients, inherently low mineral stocks in these soils, and the harsh climate of the interior plains and 

plateaus aggravate the consequences of nutrient depletion. The estimated net annual losses of nutrients vary 

considerably by sub-region: 384 800 metric tonnes for North Africa, 110 900 metric tonnes for South Africa, 

and 7 629 900 metric tonnes for Sub-Saharan Africa as a whole. This represents a total loss of US$ 1.5 billion 

per year in terms of the cost of nutrients as fertilizers. 


261 


N P K 

Country 

1982-84 2000 1982-84 2000 1982-84 2000 

(kg ha -1 yr -1 ) 

Benin 

-1 

4 

-1 

6 

-1 

-2 -9 

-1 

1 

Botswana 0 -2 1 0 0 -2 

Cameroon -20 -21 -2 -2 

-1 

2 

-1 

3 

Ethiopia -41 -47 -6 -7 -26 -32 

Ghana -30 -35 -3 -4 

-1 

7 -20 

Kenya -42 -46 -3 

-1 

-29 -36 

Malawi -68 -67 

-1 

0 

-1 

0 -44 -48 

Mali -8 

-1 

1 

-1 

-2 -7 

-1 

0 

Nigeria -34 -37 -4 -4 -24 -31 

Rwanda -54 -60 -9 

-1 

1 -47 -61 

Senegal 

-1 

2 

-1 

6 -2 -2 

-1 

0 

-1 

4 

United Republic of Tanzania -27 -32 -4 -5 

-1 

8 -21 

Zimbabwe -31 -27 -2 2 -22 -26 

Table 9.3 Estimated nutrient balance in some SSA countries in 1982-84 and forecasts for 2000. Surce: Stoorvogel and Smaling, 1990; 

Roy et al., 2003. 

More nitrogen and potassium than phosphorus get depleted from African soils. Nitrogen and potassium 

losses primarily arise from leaching and soil erosion. These soil problems result mainly from continuous 

cropping of cereals without rotation with legumes, inappropriate soil conservation practices, and inadequate 

amounts of fertilizer use. Among West African countries, Guinea Bissau and Nigeria experience the highest 

annual losses of nitrogen and potassium. Nitrogen loss in East Africa is highest in Burundi, Ethiopia, Malawi, 

Rwanda, and Uganda, and phosphorus loss is highest in Burundi, Malawi, and Rwanda (IFPRI, 1999). 

Responses to nutrient decline 

The negative nutrient balances clearly indicate that not enough nutrients are being applied in most areas 

(Bationo et al., 2012). Annual application of nutrients in SSA averages about 10 kg of NPK per ha. Fertilizer tends 

to be used mostly on cash and plantation crops because of the higher profitability of fertilizer application in 

the production of cash crops. Food crops receive less fertilizer because of unfavourable crop/fertilizer price 

ratios and financial constraints faced by farmers. In addition, food crops are only partly commercialized. 

To maintain current average levels of crop production without depleting soil nutrients, Africa as a whole 

(including North Africa) would require approximately 11.7 million metric tonnes of NPK each year, roughly 

three times more than the continent currently uses (3.6 million metric tonnes) (Henao and Baanante, 1999). 

Of this quantity, Sub-Saharan Africa would need by far the largest proportion (76 percent) because the current 

average level of fertilizer use is so low. Total nutrient requirements per ha per year range from Botswana’s 

24.5 kg ha -1 NPK (a figure 350 percent above current usage) to Reunion’s 437.3 kg ha -1 NPK (about 20 NPK per 

ha less than the country consumes). Burkina Faso would have to increase its NPK consumption more than 11 

times to maintain crop production levels without depleting nutrients and Swaziland would have to double 

its consumption. Estimated average use for SSA as a whole would have to increase about 4 times to meet 

nutrient needs at the current level of production. Generally, more nitrogen is required than potassium, and 

more potassium than phosphorus. 


262 


9.5 | Case studies 

9.5.1 | Senegal 

Introduction 

The main objective of the national land resources assessment undertaken in Senegal between 2000 

and 2010 was to identify for each land use system the status and trends of land degradation and the major 

sustainable land management interventions present in the country. This assessment used national and local 

technical expertise, including that of the land users themselves. The findings have been reported in several 

documents, maps and web-sites (Ndiaye and Dieng, 2013). 

The methodology was based on the premise that land degradation is largely driven by the way people use 

the environment in which they live. The level of degradation or sustainable use of a given land resource depends 

to a great extent on the needs and objectives of the land user, which are limited by technical knowledge 

and level of access to production factors (capital, labour etc.). Choices about land use take place within an 

integrated production system (Jouve, 1992). Consequently, defining the units in which both degradation and 

sustainable land management are to be described requires the identification of areas with similar geographic 

characteristics and then the mapping of the different production or land use systems. In Senegal this mapping 

was carried out using the ‘Framework for characterization and mapping of agricultural land use’ (George and 

Petri, 2006). 

In Senegal the following major land use systems were identified and characterized: 

1. Aquaculture and fishing, which takes place in areas covered with mangrove and other aquatic 

vegetation that are regularly flooded. 

2. Rainfed subsistence agriculture, which is characterized by the absence of livestock and minimal use of 

inputs. 

3. Agropastoral systems, characterized by a significant presence of rainfed agriculture but with greater 

levels of livestock activity. These systems are located in areas that receive between 400 and 700 mm 

of rainfall. 

4. Riverbank agriculture, characterized by the use of receding floodwaters to produce crops. 

5. Irrigated agriculture, characterized by intensive management and relatively high use of inputs. 

6. Forest based systems that exploit trees for timber. 

7. Conservation areas that are protected to preserve biodiversity 

8. Peri-urban agriculture, characterised by a mix of activities aimed at producing high-value products 

close to urban markets. 

9. Nomadic grazing, which takes place in the driest areas of the country and is characterized by shifting 

livestock and no permanent agriculture. 

The distribution of these land use systems and their extent within the country are given in Figures 9.5 

and 9.6 respectively. 

The national assessment of land degradation and sustainable management was carried out using available 

hard data and a questionnaire developed by FAO in collaboration with WOCAT (Liniger et al., 2011). The 

evaluation describes the actual situation and assesses the trends of land change over the last ten years. The 

method used local observations and measurements and expert opinion and covered the whole of Senegal. 

It has achieved the identification and characterization of land change in terms of degradation types, their 


263 


extent, degree, level, trend, causes and impacts on ecosystem services. Each of these parameters has been 

mapped and examples are given in Figure 9.7 (extent of dominant degradation type) and Figure 9.8 (rate of 

change of degradation). All information collected has been captured in a national database and analysed 

statistically (SOW-VU, 2010). 

Figure 9.5 Major land use systems in Senegal. Source: FAO, 2010. 

Figure 9.6 Proportional extent of major land use systems in the 

Senegal. Source: Ndiaye and Dieng, 2013. 


264 


Figure 9.7 Extent of dominant degradation type in Senegal. Source: FAO, 2010. 

Land degradation in Senegal - Average rate of degradation 

Legend 

Rate of increasing degradation 

Senegal 

Mauritania 

0.00 no change 

0.01 

0.5 

1 slowly increasing 

1.5 

2 moderately increasing 

2.5 

no data 

2.5 

Mali 

Atlantic 

Ocean 

Gambia 

Guinea-Bissau 

Guinea 

0 25 50 100 km 

Figure 9.8 Average rate of degradation in Senegal. Source: FAO, 2010. 


265 


Some examples of analysis of the data show the value of the national assessment: 

1. The analysis of the socio-economic drivers of land degradation in the country showed that poverty and 

population pressure are the main drivers, while governance and education are also significant. Land 

tenure and conflict situations were reported to be of minor importance as drivers of land degradation 

in Senegal. 

2. It is population pressure and poverty which lead in a majority of cases to deforestation and overgrazing 

and which, together with lack of access to extension services, lead to unsustainable soil and crop 

management. Urbanisation and mining are minor pressures in the country. 

3. Impacts of land degradation on ecosystem goods and services fell mainly on the productive services 

(affecting 15 percent of the area), while impacts on ecological services, in particular on biodiversity, 

were slightly less (13 percent). The influence on the socio-cultural provisioning services was the 

smallest, affecting only 6 percent of the area. 

4. 

Further field and socio-economic studies were undertaken at the local level, both in areas that were 

considered ‘hotspots’ for degradation and in ‘bright spots’ where degradation was less prevalent and 

sustainable management was practiced. 

The analysis of results in these local areas is illustrated in Figure 9.9 which gives the impact of degradation 

on the various ecosystem services. There is a major impact on the net returns of the farmers in all areas, but 

there are also important differences according to the different situations in each zone. 

Figure 9.9 Impact of degradation on ecosystem 

services in the local study areas in Senegal. 

Source: Ndiaye and Dieng, 2013. 

Responses have included measures implemented by the government, NGOs, the communities themselves 

and local producers. The principal responses were: assisted natural regeneration, agro-forestry, application of 

organic amendments, introduction or extension of fallow periods, composting, and using a millet/groundnut 

rotation. Most of these responses have proved to be efficient, but their adoption by land users has been slow, 

affected by lack of information and by economic and/or political constraints. 

9.5.2 | South Africa 

Of South Africa’s total area of 123.4 million ha, arable land accounts for only 11 -1 5 percent. Of this arable 

land, only about a quarter is high potential. This high potential land is thus a critical resource which needs to 

be protected. South Africa’s soils are diverse and complex as a result of varied soil formation and weathering 

processes. The largest proportion (81 percent) are slightly weathered and calcareous soils. More than 30 

percent of soils are sandy (e.g. less than 10 percent clay content) and almost 60 percent of soils have low 

organic matter content (Scotney, Volschenk and Van Heerden, 1990). The most important soil limitations 

are shallow depth, extremes of texture, rockiness, severe wetness and high erosion hazard. In terms of soil 


266 


management, it is important to note that agriculture in South Africa has a dualistic nature, with a welldeveloped 

commercial sector on the one hand, and a predominantly subsistence or small-scale agricultural 

sector in communal areas on the other. 

The first, and to date only, nationwide study of soil distribution was done by the Land Type Survey Staff 

(2003) from 1970 to 2003. The study delineated areas known as ‘land types’ at 1:250 000 scale - land types 

were defined as areas displaying a marked degree of uniformity in terms of terrain form, soil pattern and 

climate. The study included an in-depth analysis of a number of soil profiles, termed modal profiles, selected 

to represent the range of soils encountered during the survey. Soils from 2 380 profiles across the country 

were described and analysed for morphological and chemical data and classified according to the binomial 

classification system developed for South Africa (MacVicar et al., 1977). Each land type includes a collection of 

soils and their relative distribution in terms of area per landscape position, as well as their characteristics in 

terms of physical and chemical properties. The resultant national map of broad soil patterns is shown in Figure 

9.10. 


267 


Figure 9.10 Broad soil patterns of South Africa. Source: Land Type Survey Staff, 2003. 


268 


From 2006, several further national studies were conducted to assess the status of soils, land use trends, 

land degradation and sustainable land management implementation in the country. Results are summarized 

in this section. There is much evidence of mismanagement of soil resources which has led to widespread 

erosion by both wind and water, loss of soil fertility, compaction and acidification. 

National land degradation assessment 

Since land use is considered the single most important driver of land (and soil) degradation, land degradation 

assessments were conducted based on land use categories. For this purpose, a national stratification map 

was developed for South Africa based on amendments to the Land Use System Approach as described by 

Nachtergaele and Petri (2008) as well as by Pretorius (2009). On this basis, the stratification map adopted the 

following 18 land use categories: 

• Desert 

• Azonal vegetation 

• Savanna 

• Forest 

• Grassland 

• Nama-Karoo 

• Indian Ocean Coastal Belt 

• Succulent Karoo 

• Fynbos 

• Albany Thicket 

• Open Water 

• Urban 

• Cultivated – commercial – rain-fed 

• Cultivated – irrigated 

• Cultivated – subsistence – rain-fed 

• Plantations 

• Mines 

• Protected areas 

By integrating the local municipality boundaries with those of land use, a total of 2 447 unique units were 

derived for further assessment, as illustrated in Figure 9.11. Land degradation and sustainable land management 

implementation were then assessed in each of the 2 447 mapping units. The assessment was carried out from 

2008 to 2010 and was based on a participatory approach as part of a Land Degradation Assessment in Drylands 

(LADA) project. The approach relied strongly on the inputs from a range of experienced contributing specialists 

and land users who were conversant with the areas to be assessed. Data capturing was done through a series 

of 33 Participatory Expert Assessment (PEA) Workshops throughout the country, involving 728 contributing 

specialists (Wiese, Lindeque and Villiers, 2011). 

In terms of soil, the main forms of degradation at national level were soil erosion by water, biological 

degradation and chemical soil degradation. For soil erosion, sheet and gully erosion were considered the 

most serious threats, with river or stream bank erosion and off-site sedimentation considered less critical. 

Soil acidification and salinization were also highlighted, although their occurrence was more localized and 

area-specific. 


269 


Figure 9.11 The national stratification used for land degradation assessment in South Africa, incorporating local municipality 

boundaries with 18 land use classes. Source: Pretorius, 2009. 

Erosion assessment 

A separate spatial study was conducted on the extent of gully erosion in South Africa (Le Roux et al., 2008). 

The study also assessed national erosion potential in terms of soils, climate and topography (Le Roux, 2012). 

The assessment of water erosion susceptibility indicated that around 20 percent (26 million ha) of the country 

is classified as having a moderate to severe erosion risk (mainly based on sheet-rill erosion). The affected 

areas are concentrated in the south-eastern and north-eastern interior, mainly in the Eastern Cape, KwaZulu- 

Natal, Mpumalanga and Limpopo Provinces. All of these areas are characterized by a combination of high 

(often intense) rainfall, duplex soils derived from sodium-rich parent materials, and steep slopes (see Figure 

9.12). These natural conditions are often exacerbated by poor land use practices, such as incorrect cultivation 

methods, overgrazing by livestock and high population density. Under such circumstances, potential soil loss 

can easily be in the ‘Very High’ class of more than 50 tonnes ha -1 yr -1 (Le Roux et al., 2006). 


270 


Figure 9.12 Actual water erosion prediction map of South Africa. Source: Le Roux et al., 2012. 

The erosion process starts when the vegetation cover is disturbed or removed, allowing the rainfall to 

impact directly on bare soil. If measures to restrict surface run-off are not put in place, the effect is generally 

two-fold: firstly, the water flowing on the soil surface removes a significant amount of topsoil (‘sediment’), 

especially on steeper slopes; and secondly, the duplex nature of the soils (sandy topsoil abruptly overlying a 

structured clay subsoil) results in the formation of a surface seal. As a result, very little water is actually able to 

infiltrate the soil. Research in South Africa (Levy, 1988; Rapp, 1998; Bloem, 1992) indicated that exchangeable 

sodium percentage (ESP) values play an important role in erosion risk, with problematic values being over 12, 

although values as low as 5 or 6 (Bloem and Laker, 1994; Laker and D’Huyvetter, 1988) can also cause erosion 

under poor land use conditions. 

Combating soil erosion by water remains a huge challenge in many affected areas of the country due to 

a combination of lack of resources, poor knowledge or awareness and poor infrastructure, mainly roads. 

The challenges of treating erosion and the more difficult task of rehabilitating large areas of land, combined 

with the off-site effects such as silting up of dams, together pose one of the most serious soil management 

challenges in South Africa today. 

Soil nutrient depletion, acidity and organic matter 

Although soil erosion by water was confirmed as the main soil degradation type in the country, there are 

areas in South Africa affected by wind erosion, nutrient depletion, loss of organic matter, soil acidity, salinity 

and sodicity as well as pollution from mining and industrial sources. Desktop assessments of soil nutrient 

depletion, acidity and organic matter in South Africa were conducted during 2007-2008 (Beukes, Stronkhorst 

and Jezile, 2008a,b; Du Preez et al., 2010; Du Preez et al., 2011a,b; Rantoa, Du Preez and Van Huyssteen, 2009). 


271 


Nutrient depletion and acidity 

A multitude of soil nutrient and acidity studies have been conducted over time in South Africa (Bierman, 

2001; Bloem, 2002; Buhmann, Beukes and Turner, 2006; Conradie, 1994; Eweg, 2004; Farina, Manson and 

Johnston, 1993; Mandiringana et al., 2005; Meyer et al., 1998; Miles and Manson, 2000; Thibaud, 2005). These 

studies included extensive reviews of international and national documentation, interviews with experts 

from various national and provincial institutions, and processing of data available from a number of national 

databases and soil analytical laboratories. Detailed results have been reported for each of the nine provinces 

in South Africa, but only a national summary is presented here (Beukes, Stronkhorst and Jezile, 2008a,b) with 

a focus on the agricultural sector. 

The impact of the dualistic nature of South African agriculture on soil nutrient depletion was clearly 

evident, with soils from the resource-poor/small-scale/upcoming farmers generally being acidified, severely 

P depleted, and N, K, Ca and Mg deficient (Manson, 1996; Beukes, Stronkhorst and Jezile, 2008b). Within this 

group, two sub-groups can be distinguished, as these farmers produce crops at two levels. The first sub-group 

is the home garden where relatively high fertility levels are evident. This is mainly because these gardens are 

located next to the homesteads and are therefore easier to manage. The second sub-group consists of crop 

fields which are larger and further from the homesteads. As a result, these fields are less secure in terms of 

livestock access and there are transport constraints. In addition, most smallholder farmers are risk-averse due 

to their limited resources. These fields are therefore generally severely nutrient depleted, especially in terms of 

P and K deficiencies, while N, Mg and Ca deficiencies are also often noted. 

By contrast, commercial agriculture operates on a much larger scale and higher levels of management and 

inputs are maintained on these farms to ensure higher productivity. This is especially the case in the sugar, vine 

and fruit farming sectors due to higher costs for crop establishment and maintenance. Soils in the commercial 

sector generally exhibit P deficiency as the main nutrient concern, with K deficiency also occurring in many 

areas. Commercial pastures may have fewer deficiencies: for example in KwaZulu-Natal, P deficiency is almost 

negligible and K, Ca and Mg appear well supplied (Beukes, Stronkhorst and Jezile, 2008b). 

Naturally occurring acid soils are generally associated with high rainfall areas and certain geological 

materials which, in South Africa, are located in the western and southern Cape coastal belts, KwaZulu-Natal, 

Mpumalanga and Limpopo Province (see Figure 9.13). The extent of anthropogenic soil acidity in the country is 

not easy to estimate, but general trends can be observed. In the winter rainfall region, approximately 560 000 

ha is under cultivation, and on 60 percent of this area soils indicate problems with acidity. In Kwa-Zulu-Natal, 

roughly 35 percent of the more than 660 000 ha of cultivated soils have acid saturation values above 15 

percent. In the summer rainfall area west of the Drakensberg, 37 percent of topsoils in the cropped area are 

acidified (Beukes, 1995). 


272 


Figure 9.13 Topsoil pH derived from undisturbed (natural) soils. Source: Beukes, Stronkhorst and Jezile, 2008a. 

Soil organic matter 

Although data on soil organic carbon in South Africa are limited, fragmented and uncoordinated, general 

trends of SOC content can be derived from a range of studies conducted (Barnard et al., 2000; McKean, 1993, 

Du Toit and Du Preez, 1993, Le Roux et al., 2005; Mills and Fey, 2004; Prinsloo, Willshire and Du Proez, 1990; 

Van Antwerpen and Meyer, 1996; Birru, 2002; Du Preez, Mnkeni and Van Huyssteen, 2010, 2011a, b). A review of 

SOC research estimated that approximately 58 percent of South African soils contain < 0.5 percent organic C, 

38 percent contain 0.5–2 percent organic C and 4 percent have > 2 percent organic C. These organic C contents 

vary greatly as a function of soil types, climate, vegetation, topography and soil texture, and are greatly 

influenced by management practices which result in organic C losses such as overgrazing, high levels of soil 

disturbance during cultivation, and the use of fire in rangeland management. Soil organic matter losses were 

generally associated with dryland cropping, but were less prevalent in irrigated agriculture. Increasing SOM 

is a slow process, but it has been achieved by implementing zero/minimum tillage, by mulching and through 

reversion of cropland to perennial pastures. Increases have mainly occurred in the upper 300 mm of soil, and 

in most instances, have been restricted to the upper 50 mm of soil. 

Loss of SOM has been found to result in lower nitrogen and sulphur reserves, but not necessarily in lower 

phosphorus reserves. Loss of SOM also coincided with changes in the composition of amino sugars, amino 

acids and lignin. It further resulted in a decline of water stable aggregates which are essential in the prevention 

of soil erosion. 

Rantoa, Du Preez and Van Huyssteen, (2009) used data from the approximately 2 200 modal profiles from 

the land type survey to estimate organic carbon stocks in South African soils with reference to master horizons, 

diagnostic horizons, soil forms and land cover classes. In summary, the average organic carbon content in the 


273 


master horizons ranges from 16 percent in the O horizon to 0.3 percent in the C horizons. In the diagnostic 

horizons, the highest average organic C in topsoils ranged from 21 percent in the O horizons to 1.4 percent in 

the orthic A horizons. In the diagnostic subsoil horizons, however, values ranged from 1.2 percent in podzol B 

to 0.2 percent in the dorbank horizons. 

Land Cover Change Assessment 

Land use trends give an indication of land conversion from one land use to another, which directly affects 

soil use properties as a function of management. A study of land-cover change was conducted in 2010 

based on land-cover data from 1994/1995, 2000 and 2005 employing a cost-effective approach which used 

Earth Observation data. The study was based on changes in five land-cover classes: urban, mining, forestry, 

cultivation, and other. These five classes are defined in Table 9.4 (Schoeman et al., 2010). 

Land-cover class 

Urban 

Mining 

Forestry and plantations 

Cultivation 

Other 

Class definition 

Human settlements, both rural and urban 

Areas covered by mining and related mining activities 

(also includes mine dumps) 

All forestry and plantations including woodlots and clear fell areas 

(excludes indigenous natural forests) 

All areas used for agricultural activities, including old fields 

and subsistence agriculture 

All other areas not covered by those listed above 

Table 9.4 Definitions of the five land-cover classes on which the land-cover change study was based. Source: Schoeman et al., 2010. 

The land-cover change results (Figure 9.14) indicated that at national level there was a total increase of 

1.2 percent in transformed land, specifically associated with Urban, Cultivation, Forestry & Plantation and 

Mining. This represents an increase from 14.5 percent transformed land in 1994 to 15.7 percent in 2005 across 

South Africa. On a national basis the areas of Urban, Forestry & Plantation, and Mining have all increased 

over the 10-year period, whereas Cultivation areas have decreased. Urban has increased from 0.8 percent to 

2 percent, Forestry & Plantation from 1.2 percent to 1.6 percent, and Mining from 0.1 percent to 0.2 percent, 

while Cultivation has decreased from 12.4 percent to 11.9 percent. The spatial patterns do, however, vary 

geographically across provinces in South Africa. 

The increase in urban and mining areas are the biggest concern in terms of soil conservation and future 

use since urban development involves soil sealing which irreversibly removes soils from other land uses, while 

mining results in serious chemical and physical soil degradation which can only be restored to a limited extent. 

For this reason, it is essential that soil suitability and potential for agricultural and environmental purposes 

be assessed in order to ensure that the high potential and environmentally important soils are reserved and 

conserved for food production purposes. 


274 


Figure 9.14 Change in land-cover between 1994 and 2005 as part of the Five Class Land-cover of South Africa after logical corrections. 

Source: Schoeman et al., 2010. 

9.6 | Summary of conclusions and recommendations 

Based on the above finding, an assessment is made of the status and trend of the ten soil threats in order 

of importance for the region. At the same time an indication is given of the reliability of these estimates (Table 

9.5). 

Soil degradation is considered one of the root causes of stagnating or declining agricultural productivity in 

SSA. Unless soil degradation can be controlled, many parts of the continent are expected to suffer increasingly 

from food insecurity. If this decline in the productivity of Africa’s soil resources continues, the consequences 

will be severe, not only for the economies of individual countries, but also for the welfare of the millions of 

rural households dependent on agriculture for meeting their livelihood needs. 

There is an urgent need for proactive interventions to arrest and reverse soil degradation. Rehabilitation of 

degraded land and conservation of those not yet degraded is the most desirable step for every country in the 

region, but this can only be achieved if the characteristics of the soil resources are well defined and quantified 

and soil monitoring systems established in every country. 


275 


Threat 

to soil 

function 


Organic 

carbon 

change 

Nutrient 

imbalance 

Loss of soil 

biodiversity 

Soil 

acidification 

Summary 

Soil erosion constitutes 

>80% of land degradation 

in SSA, affecting 

about 22% of agricultural 

land and all countries in 

the region. 

The majority of causes 

related to the exposure 

of the bare soil surface by 

cultivation, deforestation 

overgrazing and drought. 

The replacement of 

the natural vegetation 

reduces nearly always 

the soil carbon level. 

Further carbon release 

from the soil is caused by 

complete crop removal 

from farmlands, the 

high rate of organic 

mater decomposition by 

microbial decomposition 

accentuated by high soil 

temperature and termite 

activates in parts of SSA. 

Nutrient imbalance, which 

is generally manifested by 

the deficiency 

of key essential nutrients 

is mainly due to the fact 

that fertilization has not 

been soil and crop specific, 

farmers are unable to pay 

the price for fertilizers 

and the inability to 

follow the rates that are 

recommended. Nearly all 

countries in the region 

show a negative nutrient 

balance. 

SSA suffers the world’s 

highest annual 

deforestation rate. The 

areas most affected are 

the in the moist areas 

of West Africa and the 

highland forests of the 

Horn of Africa. Cultivation, 

introduction of new 

species, oil exploration 

and pollution reduce 

the population of soil 

organisms thus reducing 

faunal and microbial 

activities. 

Over 25% of soils in Africa 

are acidic. Most of these 

occur in the 

wetter parts of the 

continent. In South Africa 

it poses as a serious 

chemical problem and 

the greatest productionlimiting 

factor. 

Condition and Trend 

Confidence 

Very poor Poor Fair Good Very good In condition In trend 


276 


Waterlogging 

Compaction 

Soil sealing 

and land take 

Soil pollution 

Most waterlogging threats 

are due to rise in water 

table due to 

poor infiltration/ 

drainage or occurrence 

of impervious layer in the 

subsoil. Waterlogging 

generally reduces crop 

productivity, but in paddy 

fields is deliberate and 

beneficial. 

The major cause of 

compaction is pressure 

on the soil from heavy 

machinery. It is more 

serious in forested regions 

where land clearing (and 

even other cultivation 

activities) cannot be done 

without mechanization. 

These constitute problems 

mainly in peri-urban 

agriculture and valley 

sites used for dry season 

vegetable production. 

Soil contamination by 

chemicals (fertilizers, 

petroleum products, 

pesticides, herbicides, 

mining) has affected 

agricultural productivity 

and other ecosystem 

services negatively. Nigeria 

and South Africa are the 

most affected. 

Table 9.5 Summary of soil threats status, trends and uncertainties in Africa South of the Sahara. 


277 


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of Soil Change in Asia 

Regional Coordinator/Author: Kazuyuki Yagi (ITPS/Japan) 

Contributing Authors: Fahmuddin Agus (Indonesia), Tomohito Arao (Japan), Milkha S. Aulakh (ITPS/India), 

Zhaohai Bai (China), Rodel Carating (The Philippines), Kangho Jung (South Korea), Atsunobu Kadono (Japan), 

Masayuki Kawahigashi (Japan), Seung Heon Lee (South Korea), Lin Ma (China), G.P. Obi Reddy (India), G. S. 

Sidhu (India), Yusuke Takata (Japan), Tran Minh Tien (Vietnam), Renkou Xu (China), Xiaoyuan Yan (China), 

Kazunari Yokoyama (Japan), Fusuo Zhang (China), Dongme i Zhou (China). 


287 

in Asia


This chapter describes the status of soil resources in the member countries of the Asian Soil Partnership 

(ASP), which includes East Asia (five countries: China, Democratic People’s Republic of Korea, Japan, Mongolia, 

and Republic of Korea), Southeast Asia (11 countries: Brunei Darussalam, Cambodia, Indonesia, Lao People’s 

Democratic Republic, Malaysia, Myanmar, Philippines, Singapore, Thailand, Timor Leste, and Viet Nam), 

and South Asia (eight countries: Afghanistan, Bangladesh, Bhutan, India, Maldives, Nepal, Pakistan, and Sri 

Lanka). 

Asia, the Earth's largest and most populous continent, is located largely in the eastern and northern 

hemispheres. The region, which consists of the above-mentioned ASP countries, covers 4.1 percent of the 

Earth's total surface area and comprises 16 percent of its land area. With approximately 3.9 billion people, Asia 

hosts 55 percent of the world's population. Population density is high – averaging 1.87 persons ha -1 – compared 

to the world average of 0.54 person ha -1 . Like most areas of the world, Asia has experienced rapid population 

growth rate in the modern era. In the twentieth century, Asia's population nearly quadrupled, as did the world 

population. 

In general, Asia enjoys a warm and seasonally humid climate and is well-endowed with natural resources 

for agriculture. The unique combination of the monsoon climate and the exceptionally large lowland area has 

made Asia the rice basket of the world (Kyuma, 2004). Sustained high levels of staple food production have 

enabled Asian countries to support a large population within a limited area of arable land. However, recently, 

Asian countries have faced rapid changes in both socio-economic and natural factors and these have had major 

impacts on agro-environments in the region. In particular, rapid economic development and urbanization are 

changing land management systems in many countries, and climate change has emerged as a significant 

source of risks. These changes are having major impacts on the status of soil resources in the region. 

10.2. Stratification of the region 

10.2. 1 | Climate and agro-ecology 

The map of Asia shows many vast rivers with large alluvial plains and deltas. Major rivers include the Yellow, 

Yangtze, Mekong, Chao Phraya, Irrawaddy, Ganges, Brahmaputra, and Indus. The Himalayan mountain range 

runs for more than 2 400 km, separating the Indian subcontinent from the rest of Asia. Many of Asia's major 

rivers have their source in the Himalayas and adjacent plateaus. These rivers have created vast areas of fertile 

land suitable for farming. On the east and southeast shores of the continent lie a number of islands chains or 

island arcs, many characterized by mountainous landscape with volcanic activities. 

Most areas of the Asian region are strongly influenced by the monsoon, a seasonal reversing wind 

accompanied by corresponding changes in precipitation. For this reason, the region is often called ‘Monsoon 

Asia’. The Asian monsoon is a highly significant component of the global climate system. It has a huge influence 

on how people live and on their livelihoods, and it provides water resources throughout the region (Salinger 

et al., 2014). The East Asian monsoon carries moist air from the Pacific Ocean and the Indian Ocean to East 

and Southeast Asia in summer. In winter, it reverses and carries cold, dry air from the Eurasian Continent. In 

south Asia, winds blow from June to September from a south-westerly direction from the Indian Ocean onto 

the Indian landmass, bringing rain to most parts of the subcontinent. Subsequently, from around October, 

the winds reverse direction and start blowing from a north-easterly direction, from the subcontinent onto 

the Indian Ocean. These winds carry less moisture and bring rain to only limited parts of India. Dry areas 

predominate in parts of the north-western interior where the influence of oceanic winds is less. This wide 

diversity of climate is a major factor in the stratification of the region into different agro-ecological zones 

(Figure 10.1) 


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in Asia

Figure 10.1 Length of the available growing period in Asia (in days yr -1 ). Source: Fischer et al., 2012. 

A distinctive feature of monsoon Asia is its very high share of the world’s area and production of rice (Kyuma, 

2004). Some 90 percent of the global acreage and output of rice are concentrated in the region, earning it the 

title of ‘rice granary’ of the world. This dominance of rice cultivation is due to several factors, notably high 

precipitation and temperatures and the existence of extensive lowlands suitabe for paddy prduction. The vast 

expanse of lowlands is a unique feature of the region, resulting from a combination of geological instability and 

the high precipitation. Rice cultivation originally emerged as an adaptation to extensively inundated lowlands, 

but with time it was expanded to areas that could support rice only with irrigation. High productivity and high 

sustainability are the outstanding advantages of rice cultivation. By contrast, upland cultivation of dryland 

crops in Monsoon Asia is handicapped by low soil fertility and high susceptibility to soil erosion. 

10.2.2 | Previous regional soil assessments 

Based on the report by Oldeman (1991), the GLASOD project estimated that human-induced soil 

degradation in Asia region (including non-ASP west Asian countries) accounted for 31 percent of the inhabited 

land area, the highest share of any of the global regions. Soils in Asia were found to have been degraded by 

several factors: water erosion (59 percent), wind erosion (30 percent), chemical degradation (10 percent) and 

physical degradation (2 percent). Deforestation was identified as the most dominant causative factor for soil 

degradation, followed by agricultural activities and overgrazing. In most Asian countries the mining of soil 

nutrients is causing decline of average crop yields. Fertile soil is washed away by the erosive forces of water, or 

blown away by wind. This so-called first generation of environmental problems is leading not only to negative 

nutrient balances but also to habitat destruction and loss of biodiversity (Oldeman, 2000). 


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in Asia

Following GLASOD, the need for more detailed and more country-specific degradation assessments 

became apparent. In 1993, the members of the Asian Network on Problem Soils recommended the preparation 

of a qualitative assessment for South and Southeast Asia at a scale of 1:5 million. This recommendation was 

acknowledged by FAO and UNEP. FAO assigned ISRIC to prepare a new physiographic map and database 

at 1:5 million scale. UNEP prepared and implemented the Assessment of the Status of Human-Induced Soil 

Degradation in South and Southeast Asia (ASSOD). Sixteen national institutions for natural resources in the 

region collaborated on the project under the coordination of ISRIC. The ASSOD project published revised subregional 

Guidelines for General Assessment of the Status of Human-Induced Soil Degradation and produced 

regional maps on the status of human-induced soil degradation at a scale of 1:5M together with digitized 

version of the map (van Lynden and Oldeman, 1997). 

The different soil degradation types inventoried by ASSOD are described below and shown in Figure 10.2: 

• Water erosion: Water erosion covers 21 percent of the total land area in the region (or 46 percent of the 

total degraded area). It is predominant in large parts of China (>180 million ha) except for the northern 

parts, on the Indian subcontinent (>90 million ha) and in the sloping parts of Indochina (40 million ha), 

the Philippines (10 million ha) and Indonesia (22.5 million ha). In relative terms, as a percentage of the 

total country area, moderate to extreme water erosion is particularly important in India (10 percent), 

the Philippines (38 percent), Pakistan (12.5 percent), Thailand (15 percent) and Vietnam (10 percent). 

• Wind erosion: Wind erosion (9 percent of the total area, 20 percent of all degradation) is concentrated 

mainly in the most western and northern arid and semi-arid desert regions of Pakistan (>9 million ha 

on-site and >2 million ha off-site), India (20 million ha on-site, 3.6 million ha offsite) and China (>70 

million ha on-site, >8.5 million ha off-site). Although large parts of these regions may be considered 

deserts, some human-induced wind erosion was also reported. 

• Chemical deterioration: Chemical deterioration is distributed in patches, probably also partly due to 

different perceptions of this type of degradation. About 11 percent of the total area (or 24 percent of the 

degraded area) is affected by some kind of chemical deterioration. High relative extents of chemical 

deterioration (>30 percent of total country area) can be observed in Bangladesh, Cambodia, Sri Lanka, 

Malaysia, Pakistan and Thailand, generally with negligible to light impact. 

• Physical deterioration: Occurrence of physical deterioration (affecting about 4 percent of the total 

area or 9 percent of the total degraded area) is even more dispersed and infrequent than chemical 

deterioration. Waterlogging and aridification are the main subtypes, in particular in Bangladesh, 

China, India and Pakistan. Compaction or crusting/sealing is relatively unimportant except in Thailand 

and the Philippines, although they occur in most countries. Waterlogging and compaction as a result 

of paddy cultivation are not considered as degradation. Loss of productive function as a result of 

urbanisation, industrialisation and infrastructure has been indicated for only a few countries although 

this phenomenon is on the rise. 


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in Asia

Figure 10.2 Threats to soils in the Asia region by country. 

10.3 | General threats to soils in the region 

10.3.1 | Erosion by wind and water 

Soil erosion is the most important threat to soil in Asia. Water erosion by rainfall and surface water flow is 

dominant in humid regions with torrential rains, as in South and East Asia. In drier or desert areas, wind is the 

driving force inducing soil erosion. This threat is discussed below in Section 10.4.1 in detail. 

10.3.2 | Soil organic carbon change 

Data for evaluating soil organic carbon (SOC) change in Asian countries are limited because countries do 

not generally monitor SOC stock and changes. However, data from available literature show that where there 

are increases in crop yield in croplands of East and Southeast Asia, SOC is retained. SOC has also been shown 

to accumulate in forest areas. However, in South Asia SOC is decreasing. This is because crop residues are 

widely used as fuel and fodder and are not returned to the soil. In Indonesia, three anthropogenic activities 

– deforestation, poor land management, and intensive cropping – contribute to SOC change in mineral and 

peat soils (Section 10.5.2). Throughout the region, the degradation of grassland has generally caused great 

losses of SOC stock. This threat is discussed in detail below in Section 10.4.2. 


Sources of contamination of arable land in most Asian countries include (i) parent material, (ii) mining, 

(iii) smelting, (iv) agrochemicals and sewage sludge applications, and (v) livestock manure uses (Luo et 

al., 2009). There is an urgent need to reduce hazardous chemical concentrations of Cd, As, Pb, Cu and Zn, 

especially in paddy soil and rice grains. In many regions of Southeast Asia (Bangladesh, India, China, Vietnam, 

Taiwan - Province of China, Thailand and Nepal), arsenic is naturally present in groundwater and represents 

a threat to sustainable agriculture (Smedley, 2003; Brammer and Ravenscroft, 2009). This enrichment is 


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in Asia

magnified considerably when mining takes place, as it is the case of the Ron Phibun district, in southern 

Thailand (Williams et al., 1996). Arsenic contamination particularly affects rice, as this crop not only requires 

large amounts of water but is also grown under the anaerobic conditions favoured in rice fields where As is 

mainly present in its trivalent form which is readily available to plants (Brammer and Ravenscroft, 2009). 

The agricultural rice soils of the Guandu Plain in Taiwan, Province of China are seriously contaminated with 

As and Pb (Zhuang et al., 2009; Chang et al., 1999, 2007). Cadmium pollution of paddy fields has been found 

downstream of a Zn mineralized area in Thailand (Simmons et al., 2005). China contributes around 28 percent 

of global Hg emissions, with India, Japan and the Korea Democratic Republic also among the ten countries 

that contribute the most to global Hg emissions (Li et al., 2009). 

According to the Agricultural Land Soil Pollution Prevention Law in Japan, the maximum allowable limit 

of Cd in paddy fields is set in terms of the Cd concentration in rice grains produced in the field, not the soil Cd 

concentration of the field. This is because the amount of bio-available Cd in soil is affected dramatically by 

water management practices used for rice cultivation (Asami, 1981). 

With rapid industrialization, urbanization and intensive use of farmland, China is now facing serious soil 

pollution (Ministry of Environmental Protection, the People’s Republic of China, 2014). About 19.4 percent 

of farmland has high levels of Cd, Ni and As pollution. Soil contamination has been estimated to cause a 

reduction of more than 107 tonnes of food supply annually (Wei and Chen, 2001). In 2006-2010, China’s 

Ministry of Environmental Protection and Ministry of Land and Resources jointly launched a nationwide 

investigation of soil pollution status, covering an area of 6.3 million square kilometers. In 2014, a Communiqué 

on Nationwide Soil Contamination was issued by the two ministries indicating that the overall situation of 

the soil environment in China is not encouraging. Some regions are heavily polluted. There is concern over 

the quality of farmland soils and over other soil environmental issues also caused by anthropogenic activities, 

e.g. those related to mining and industrial activities which cause atmospheric deposition, and to the use of 

livestock manures (Luo et al., 2009). Trace elements are pollutants of major concern, especially in the southern 

area of China. Currently, the evaluation of soil pollution in China is primarily based on the Environmental 

Quality Standard for Soils which was promulgated and implemented in 1995. The Standard needs further 

improvement because of its stringent limitations. At present, China’s Ministry of Agriculture is working on an 

Implementation Plan on Prevention and Control of Heavy Metal Pollution in Agricultural Products. 

Historic and current rates of intensive pesticide and fertilizer use in agricultural land and also industrial 

development have caused the accumulation of organic pollutants and heavy metals in soils of Indonesia. 

Earlier research showed high concentration of organo-chlorines in vegetables (Dibiyantoro, 1998). However, 

residue levels in foodstuffs have gradually reduced to within acceptable daily intake levels as established 

by WHO (Shoiful et al., 2012; Rahmawati et al., 2013). The tapioca industry in Java is now recognized as a 

contributor to cyanide levels which have risen above background levels in river water (Indrayatie et al., 2013). 

Mining plays an important role in the Indonesian economy but this mining, particularly artisanal and smallscale 

mining (ASGM), can have a major impact on the environment (Limbong et al., 2003; Prasetyo et al., 2010). 

During ASGM, Hg is used to recover Au from the ore during grinding. The process is inefficient and releases 

a significant amount of Hg to soil, water and the atmosphere (Limbong et al., 2003; Edinger et al., 2008). 

Tailings from Hg amalgamation are then leached with cyanide. Ultimately, the final waste, contaminated 

with metals and cyanide, is released into the environment (Veiga et al., 2009). Many ASGM operations also 

release As and Sb to the environment, although this depends on the composition of the host ore (Edinger et al., 

2008). These operations are unlicenced and illegal. Indonesia has now signed the Minamata Convention on 

Hg and has decentralized control of ASGM operations to provincial governments. Environmental protection 

has become a primary objective for government regulation. 


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in Asia


There is large area of acid soils distributed across the tropical and subtropical regions of Asia, mainly in 

Southeast Asia, parts of East Asia and parts of South Asia. Acid sulphate soils are widely distributed in the 

coastal plains of Southeast Asia and southern China. The total area of acid sulphate soils in Southeast Asia is 

7.5 million ha and there are about 112 thousand ha of these soils in China (Shamshuddin et al., 2014). The soils in 

these regions are also sensitive to external acid such as acid deposition (Hicks et al., 2008). 

In Vietnam, ferralitic, basaltic, and grey degraded soils, which cover about one third of the country, have 

strong potential for acidification and degradation because of their nature (NISF, 2012). A leading contributor 

to soil acidification in Vietnam is the unbalanced and unsuitable application of chemical fertilizers. Data 

from the International Fertilizers Association (IFA) show that from 1961 to 2012, total NPK chemical fertilizer 

use in Vietnam increased 31 times (IFA, 2012). The increase in fertilizer application and unbalanced use lead 

to inadequate presentations of acidic factors in soil solutions. Some types of fertilizers (including organic 

fertilizers) can make the soil more acidic (Nguyen, 2014). Another reason for the acidification process in 

Vietnam is the presence of sulphate soil. The area of sulphate soil in the Red River delta has increased by 7 000 

ha and the content of S in this soil has also strongly increased. In the Mekong river delta, the area has reduced 

by 261 000 ha and the content of total exchangeable cations and total dissolved salts has also reduced. 

Increase in temperature, change of precipitation, and sea level rise due to climate change may also be having 

negative effects on this process in Vietnam. 

The total area of Indonesian acid upland soils (pH


Potentially the greatest contributor to soil biodiversity loss in Asia is land use change. In China, for example, 

an assessment of land-use change across the country indicated that the largest changes were associated with 

the conversion of productive cultivated land to urban areas, thereby removing fertile land from agricultural 

production. Conversion of cultivated areas, forest and grasslands between 1996 and 2008 has been estimated 

to be 1 475 × 104 ha, 269 ×104 ha, and 536 ×104 ha, respectively (Wang et al., 2012). There have been attempts 

at land restoration to maintain soil biodiversity, such as in the coastal lands in the Jiangsu Province of China. 

Generally, there is a higher diversity of soil macrofauna in uncultivated land and forests compared to less 

diverse wheat farms and bulrush land (Baoming et al., 2014). Soil faunal diversity is significantly impacted 

by land use with a strong relationship between vegetation and macrofauna distribution and composition 

(Baoming et al., 2014). In Thailand, the forested land area has decreased by more than 50 percent in the past 

40 years (Fisher and Hirsch, 2008). Although reforestation measures can be used to restore degraded lands, 

difficulties with establishment of native trees in these areas may lead to the use of substitute trees, resulting 

in long lasting effects on soil biodiversity. For example, a study of soil microbial communities of an Acacai tree 

plantation established on degraded land in Thailand found reduced microbial activities compared to natural 

evergreen forest, suggesting that key microbes had been lost (Doi and Ranamukhaarachchi, 2013). 

Studies in India have also indicated an effect of land use change on soil organisms. The Nilgiri biosphere 

reserve in the Western Ghats of India, a global hotspot for aboveground biodiversity, has been under pressure 

because of high population density in the surrounding region (Mujeeb Rahman, Mujeeb, Varma and Sileshi, 

2012). In this area, land management had a significant effect on soil macrofauna, with larger densities and 

diversity found in the forest sites and a clear response of macro-invertebrates to land use (Rossi and Blanchart, 

2005). Interestingly, there was a high similarity in macrofauna between primary forest and disturbed forest 

plots, which indicated the potential of land restoration. A separate morphological analysis of soil invertebrate 

density at 15 different land use sites (from intensively managed agricultural systems to pristine forests) 

identified a wide range of soil faunal groups including earthworms, termites, ants, grasshoppers, crickets, 

mole crickets, bugs, coccids, cicadas, woodlice, centipedes, millipedes, and spiders (Mujeeb Rahman, 

Mujeeb, Varma and Sileshi, 2012). The natural forests had significantly higher taxonomic richness at the family 

level than soils from the annual cropping system and a significantly higher total number of individuals than 

annual crops, agroforestry and plantations. The highest richness (identified to family level) was found in the 

sites with the least anthropogenic disturbance, and the greatest diversity of earthworms, ants and termites 

(determined to species level) was found in the more complex forest ecosystems. An earlier study in the region 

recorded almost double the number of earthworm species in soils from forest compared to pasture (Blanchart 

and Julka, 1997). These results indicate a decrease in earthworm biodiversity associated with forest loss and a 

lower diversity of soil macrofauna with more intensive land use. 

In addition to anthropogenic pressures on soil biodiversity, natural disturbances such as tsunamis can 

affect soil biodiversity. Tsunami-affected areas in the Phang Nga province of Thailand had a higher proportion 

of prokaryotes (archaea and bacteria; 83.25 percent) compared to non-affected areas (72.5 percent), whereas 

the non-affected areas were more hospitable to eukaryotes (animals, plants, fungi and protists) (Somboonna 

et al., 2014). Increased occurrences of tsunamis and other natural disasters may result in losses to above and 

belowground biodiversity. 

There has not been a comprehensive analysis of threats to soil biodiversity in Asia completed to date. 

Scientific evaluation of soil biodiversity over large regions has been extremely difficult due to: (1) size of 

organisms, (2) large abundance and diversity of organisms, and (3) lack of research and gaps in the information 

collected. In Japan, the fusion of complex systems theory and computer technology has made it possible to 

employ the tools of statistical physics to develop a totally new technology for assessment of farmland soils 

(Yokoyama, 1993). The technology is used primarily for a health check on farmland soil, to measure the 

environmental affinity of production. Commercial analysis services have been started to add value to the 

products (Sakuramot, Yokoyama and Iekushi, 2010; Yokoyama and Taguchi, 2013). 


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In paddy rice cultivation and upland farming with sustainable agriculture, crop rotation and organic 

amendments have generally maintained good soil microbial diversity. However, loss of significant microbial 

diversity is notable in large-scale and intensive vegetable production areas due to: (i) continuous monoculture 

cropping, (ii) over-cultivation, (iii) intensive use of chemical fertilizers, (iv) chemical fumigation of soils to 

prevent soil-borne diseases, (v) soil contamination and non-target effects of chemical pesticides such as 

fungicides, nematicides, herbicides and insecticides, and (vi) over-application of herbicides for weed control. 

The loss of diversity in the soil is conducive to further diseases and other cropping issues from processes (ii) to 

(vi). 

Recently, the effect of poor management practices on soil biodiversity has led to concerns. As a result, 

environmental and safety-oriented consumers have tried to promote organic farming and to accelerate the 

branding of sustainably produced agricultural products based on scientific evidence. The idea is to create 

incentives for farmers to consider soil biodiversity in their farming practices. The private sector has undertaken 

evaluation of microbial diversity, demonstrating high soil microbial diversity in organically managed paddy 

fields in Taiwan - Province of China and in the Philippines, as well as in upland crops in southern China. 

However, in soils of the semi-arid areas of northern China, the loss of microbial diversity is severe. Here the 

processes mentioned –in points (i) to (vi) above may lead to soil drying, increases in salt concentrations and a 

loss of stable soil surface. These changes could potentially result in loss of soil organisms, although research 

exploring the loss of ecosystem functions and services from these soils has yet to be conducted. 


Two types of waterlogging may occur – permanent waterlogging in natural swampland, and occasional 

waterlogging in flood prone areas along the flat coastal regions and flood plains of main rivers. Constructed 

wetlands such as paddy fields are intentionally flooded as part of the management system. Waterlogging 

can have other anthropogenic causes such as poor drainage systems in settlements, industrial and urban 

development, or deforestation in upstream areas, all of which may increase the threat of water logging in 

flood prone areas. 

In the GLASOD estimate, waterlogging affects 4.6 Mha in Asia, largely in the irrigated areas of India and 

Pakistan. Waterlogging is closely linked with salinization. Since the start of large scale irrigation schemes in 

the 1930s, the progressive rise in the water table beneath irrigation areas on the Indo-Gangetic plains has 

been monitored (e.g. Ahmad and Kutcher, 1992). For India, monitoring results suggest a waterlogged area 

more than twice the GLASOD estimate. For Pakistan, four sources give total areas affected by waterlogging 

of between 1.6 and 3.7 million ha, compared with the GLASOD value of 0.96 million ha. Since the Pakistan 

country data come from at least two independent surveys, show good agreement and are believed to result 

from detailed field surveys, these country estimates are likely to be more accurate than the much lower 

GLASOD estimates. 


Harvested crops remove nitrogen, phosphorus, and other nutrients from agricultural soils. Hence, sustaining 

agricultural production requires replacement of those nutrients, whether through biological processes 

like nitrogen fixation or through the addition of animal wastes or mineral fertilizer to fields (Vitousek et al., 

2009). Balanced nutrient supply is essential for achieving high crop yields, but excessive and/or imbalanced 

nutrient input may pose risks to the environment, human health and ecosystems. Nutrient inputs in Asia vary 

considerably amongst countries, areas and farming systems. 

In some countries or regions in Asia, removal of nutrients from the soil in crop harvest appears substantially 

to exceed inputs through natural replacement or fertilizer application. For example, negative soil nutrient 

balances have been reported for each of the 15 agro-climatic regions of India (Biswas and Tewatia, 1991; 

Tandon, 1992). 


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Problems are also caused by imbalances in fertilizer application. Fertilizer use in the region is often 

dominated by nitrogen (N), relative to phosphorous (P) and potassium (K). This trend originated in the 

earlyyears of the ‘green revolution’. When fertilizers are first applied to a soil, a high response is frequently 

obtained from nitrogen. The improved crop growth depletes the soil of other nutrients; “In such systems, 

nitrogen is simply used as a shovel to mine the soil of other nutrients” (Tandon, 1992). Long-term experiments 

in India show depletion of soil P and K is higher for plots with N fertilizer, and depletion of K is higher still with 

N+P fertilizer (Tandon, 1992). The use of N fertilizer in Indonesia has been continually increasing since the late 

1960s, reaching 2.4 million Mg yr -1 in 2012. However, the consumption of P and K has not followed the same 

trend (IFADATA, 2007). This is partly due to the fact that N fertilizers are cheaper and more accessible, and 

also to the rapid crop response to N fertilization. A study on nutrient balances in rice fields under intensive 

cultivation found a positive balance for N, P, Ca, Mg and Na, but a negative balance for K and Si (Husnain, 

Masunaga and Wakatsuki, 2010). 

For secondary nutrients and micronutrients, an increasing incidence of sulphur and zinc deficiency 

is occurring in the region. Sulphur deficiency has been reported for India, Pakistan and Sri Lanka, and zinc 

deficiency for India and Pakistan (FAO/RAPA, 1992). For Bangladesh, 3.9 million ha are reported to be deficient 

in sulphur and 1.75 million ha in zinc, including areas of continuous swamp rice cultivation (Bangladesh, 

1992; Shaheed, 1992). Because of its generally alkaline soils, Pakistan is particularly liable to micronutrient 

deficiencies (Twyford, 1994). 

On the other hand, nutrient additions to many fields in some Asian countries far exceed those in the United 

States and Northern Europe, and much of the excess fertilizer is lost to the environment, degrading both air 

and water quality. This important threat to soil is described more in detail in Section 10.4.4. 


Slightly compacted soil conditions are conducive to soil productivity as they reduce soil erosion and 

maintain soil structure. Highly compacted soil is, however, a physically deteriorated condition affecting 

plant productivity under various land uses. Decrease in soil porosity that affects water content, hydraulic 

conductivity and gas permeability is a major disadvantage brought about by soil compaction. Loading of 

heavy equipment strongly compresses surface soil and/or subsoil in cropland, grassland and timber forests. 

Mechanization of cultivation and harvest in Asian countries has increased, resulting in soil erosion and soil 

compaction due to tractor loading (Zhang et al., 2006). Some studies of soils in rice/wheat cropping areas of 

India showed increases in compactness of subsurface soils as indicated by increased bulk density as high as 

1.80 g cm -3 . This was due to the use of heavy machinery in conjunction with puddling activities (Sidhu et al., 

2014; Singh, Jalota and Sharma, 2009; Aggarwal et al., 1995; Kukal and Aggarwal, 2003). Heavy machines for 

harvesting and skidding logs also compress soils considerably, and may be accompanied by rutting on the 

ground and by removal of organic matter (Kozlowski, 1999; Hattori et al., 2013). The increase in the area of 

plantation forests in Asia, which has been especially rapid in China in recent decades (FAO, 2006), has led 

to compaction of soils through the use of heavy equipment for management. Livestock trampling is also a 

major cause of surface soil compaction in grassland and hilly grazing areas (Drewry and Paton, 2000). Soil 

compaction due to heavy grazing has led to severe land degradation in the extensive pastoral steppe regions 

of Mongolia and Inner Mongolia of China (Kruemmelbein, Peth and Horn, 2008). Where soils have been 

compacted in agricultural and forested areas, water surface runoff followed by soil erosion is a major threat. 

Soils in urban green areas are commonly compressed by human traffic and by vehicles for park management. 

This results in damage to plant roots and reduces productivity (Jim, 1998a) because in urban park vegetation 

over 50 percent of root density is concentrated in the top 50cm of the soil depth (Millward, Paudel and Briggs, 

2011). Surface runoff on compacted soil generates flooding and delivers contaminants such as heavy metals 

and persistent organic pollutants into receiving water environments. 


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10.3.10 | Sealing and capping 

Sealing and capping on the soil surface is mainly required in urban areas to construct roads and buildings. 

Although the impervious surface area (ISA) covers only 0.43 percent of the global land area at present 

(Schneider, Friedl and Potere, 2009), ISA is constantly on the increase, as shown by satellite image data and 

by measures of the constant increase in urban population (Elvidge et al., 2007). More than half of the global 

population is now concentrated into cities. The Asia region has the largest ISA ratio in the world (Schneider, 

Friedl and Potere, 2009). China has the largest ISA in Asia followed by India, Indonesia, Japan and Bangladesh 

(Elvidge et al., 2007). Increase in the ISA causes environmental issues such as the formation of urban heat islands 

(Changnon, 1992), increases in surface water runoff (Booth, 1991), and reduction of carbon sequestration due 

to reduction of the forested area (Milesi et al., 2003). 

Inter-regional differences in properties of sealed soils are not well documented. However, in general, 

construction processes of sealing affect soils physically and chemically due to disturbance of surface and 

subsoil by excavation and filling, and by addition of construction materials. In Japan, very large volumes of soil 

are excavated everyyear for land levelling, and much of this soil is carried away to other sites. Soils sealed by 

construction are compressed to enhance their physical strength, making them hugely compacted structures. 

Additives such as lime, which is used to enhance the sub-base strength of a road, make soils alkaline (Jim, 

1998b). Pervious asphalt paving and inter-locking covers for light traffic roads are now recommended to drain 

rainwater by infiltration to the sub-soil. However, this infiltration treatment can make soil solution and 

drainage water alkaline. Cracks on a paved road surface due to heavy traffic may allow rainwater to infiltrate 

into the subsoil, reducing the strength of the roadbed and making the sub-soil alkaline. 

10.4 | Major threats to soils in the region 

10.4.1 | Erosion 

Soil erosion is one of the major threats to soil quality in Asia. Soil erosion is the action of exogenic processes 

such as water flow and wind to move soil from its location. The processes inducing soil erosion vary with 

climate. Asia can be divided into several climate zones: tropical and subtropical in South Asia, humid 

subtropical and temperate in East Asia, semiarid in China, and arid in Mongolia and East Asia. Most regions 

of Asia are affected by the Asian-Australian monsoon which causes dry and wet seasons. Water erosion is the 

major type of erosion in the regions of South and East Asia with alternating dry and wet seasons. On the other 

hand, wind is the crucial driving force inducing soil erosion in the drier and desert areas. 

Soil erosion by rainfall and surface water flow is generally affected by five factors: rainfall erosivity, soil 

erodibility, topography, surface coverage, and support practices. In humid regions, soil erosion is of little 

concern in well-established forests and in paddy fields. However, bare lands such as logged forests, construction 

areas and upland crop fields on slopes are exposed to a high risk of soil erosion. Annual soil loss in paddy fields 

has been generally reported in case studies to be lower than 1 tonnes ha -1 (Chen, Liu and Chen, 2012; Choi et al., 

2012; Kim et al., 2013). By contrast, soil loss from upland crop areas on slopes is much greater – for example, 38 

million tonnes ha -1 from fields in South Korea where no conservation practices were applied (Jung et al., 2005). 

In semiarid regions, soil erosion is also of concern especially for slope areas with scant vegetation. In these 

areas, several hundred mm of rainfall in the rainy season can result in massive gully erosion. For these reasons, 

soil erosion is regarded as the most important threat to soil in Asia, especially for poorly-covered lands and 

bare soil. 


297 

in Asia

Major threats of soil erosion by water are found in the hilly and mountainous landscapes of Indonesia. 

Natural conditions, anthropogenic influences on land cover and intensive land use make the steep and 

densely populated islands of Java, Sumatra and Sulawesi the most threatened areas. Approaches to coping 

with erosion problems range from engineering measures such as terracing, sediment pit construction and 

waterways improvement to vegetative measures including agroforestry approaches, contour strips and cover 

crops (Agus and Widianto, 2004). 

In the Philippines, a monsoon country with high rainfall, erosion by water is one of the major causes of 

land degradation. In the 2013 Land Degradation Assessment final report by the Bureau of Soils and Water 

Management, the estimated annual soil loss for agricultural land for the whole country based on the 2003 

Land Use System Map is about 61.8 tonnes ha -1 yr -1 . Paningbatan (1987) estimates that soil loss of about 10 

tonnes ha -1 can be considered tolerable for Philippines conditions. In the 1950s, USDA (quoted by Schertz, 1983) 

established soil loss tolerance values for the Philippines at about 5 to 12 tonnes ha -1 yr -1 at soil bulk density 

of 1 200 kg m -3 . An analysis made by the Department of Environment and Natural Resources (DENR) on the 

state of the Philippine environment showed that, overall, 75 percent of total croplands are vulnerable to 

erosion of various degrees. To counter high rates of erosion, sustainable land management practices are being 

promoted. These include the application of various soil conservation and management strategies in highland 

agriculture as well as other technologies like agroforestry and multiple-storey cropping. The Philippines is 

a member of the UN Convention to Combat Desertification (UNCCD), and the Department of Agriculture 

and other government agencies, academia and non-government organizations aggressively pursue various 

programs. These approaches seek to engage farmers as partners in development rather than treat them as 

the cause of the problem. 

In South Korea, deviation of annual precipitation was 251 mm yr -1 between 1981 and 2010 and rainfall 

erosivity ranged from 2 264 MJ mm ha -1 yr -1 hr -1 to 6 856 MJ mm ha -1 yr -1 hr -1 in the same period (Park et al, 2011). 

The national average rainfall erosivity was 4 276 mm ha -1 yr -1 with EI 30 data between 1973 and 1996 (Jung et al., 

2004) and was 4 147 MJ mm ha -1 yr -1 hr -1 with EI 60 data between 1981 and 2010 (Park et al., 2011). The variation of 

rainfall is expected to increase with climate change based on the RCP scenario (CCIC, 2014), which implies that 

the probability of extreme rainfall erosivity would also increase in future. Soil erodibility in South Korea was 

0.027 million tonnes hr MJ -1 mm -1 and ranged from 0.001 million tonnes hr MJ -1 mm -1 to 0.102 million tonnes 

hr MJ -1 mm -1 with soil series. The soil erodibility of paddy fields is the greatest at 0.036 million tonnes hr MJ -1 

mm -1 followed by upland crop fields (0.026 million tonnes hr MJ -1 mm -1 ) and forests (0.020 million tonnes hr 

MJ -1 mm -1 ) (Jung et al., 2004). 

Differences in soil erodibility between land use types are affected by geological characteristics. Forests 

are located in mountains and upland crop fields are generally placed on lower slopes below the mountains. 

This means that forests and upland crop fields have been chronically exposed to past overland flow and soil 

erosion, which has resulted in their current status with less easily-eroded particles such as silt and very fine 

sand. By contrast, paddy fields in the bottomlands are formed from sediments with greater soil erodibility. 

However, actual annual soil loss in paddy fields has been generally reported to be lower than 1 Mt ha -1 in case 

studies because paddy is protected from rainfall by the water already ponded in the field and by the ridges 

which store water inside paddy fields (Chen, Liu and Chen, 2012; Choi et al.; 2012, Kim et al., 2013). Levels of soil 

erosion are tolerable in well-managed forests and grasslands (Kitahara et al., 2000; Lee, 1994). Where grass 

was sown in spring, soil loss on grasslands was found to reach only half of that from bare soil in the firstyear. 

Losses decreased abruptly after the second year, to only 3 percent or less of those on bare soil (Jung, 1998; Jung 

and Oh, 1993). 

Soil loss from upland crop fields is much greater than from paddy fields and forests. The national average 

of soil loss in upland crop fields was 38 million tonnes ha -1 in South Korea (Jung et al., 2005). A variety of 

conservation practices have been applied to reduce soil erosion including agronomic and engineering 


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practices. On low slopes, agronomic practices such as mulching and grass strips can reduce soil loss by 80-90 

percent. Engineering practices such as terraces, channels and drop spillways can reduce soil loss on steeper 

slopes (Jo et al., 2009). Based on the Law of Water Quality Conservation and the Law of Soil Environment 

Conservation, soil erosion in South Korea has been monitored in severely eroded areas and best management 

practices have been proposed to regional governments and farmers. 


ASSOD describes soil organic carbon change as “a net decrease of available nutrients and organic matter 

in the soil”. Not all Asian countries monitor the soil organic carbon (SOC) stock and its change. Even where 

soil properties are monitored, some data may not be appropriate for aggregation at the country scale as they 

lack certain parameters such as bulk density, or the number of observations may be insufficient. However, 

data compiled from the literature are adequate to draw a rough sketch of SOC change in the region. Piao et 

al. (2012) compared three methods to estimate SOC change 1990-2009 in five East Asian countries. Using an 

inventory-remote sensing model approach, they concluded that the sub-region was a net ecosystem carbon 

sink (+0.293 Pg C yr -1 ). They found an estimated net SOC increase in the forest (+0.014 Pg C yr -1 ), shrub land 

(+0.022) and cropland (+0.022) and SOC decrease in the grassland (-0.003). 

China occupies the largest land area of the region (46 percent of the total). China reported SOC changes 

in the range -0.143 Pg C yr -1 to +0.094 Pg C yr -1 during 1980-2000. India occupies the second largest area (14 

percent of the total). In India, it is estimated that forest accumulated SOC at the rate of +0.041 Pg C yr -1 over 

the one hundred year period 1880 -1 981. The relevant data are listed in Table 10.1. Overall, there is a tendency 

towards SOC accumulation in forested areas and towards a decrease in grassland areas. To confirm these 

findings and allow analysis to guide future decisions, more detailed and comparable datasets for both SOC 

stock and change should be compiled in the region. 

Sub-region/Country 

Area 

M ha 

SOC stock 

Pg C 

SOC stock 

-1Mg C ha-1 

Mg C ha 

SOC change 

-1Pg C yr-1 

Pg C yr 

Reference 

East Asia 

sub-region 

Total 1990 - 2010 1156 0.055 Piao et al. (2012) 

Total 1980 - 2000 

871 89.6 102.9 

86.8 -0.143 

Xie et al. (2007) 

China 

Total 1981 - 2000 

760 85.8 112.9 

87.6 0.094 

Tian et al. (2011) 

Total 2000 101.9 0.029 Houghton and Hackler (2003) 

Total 0.075 Piao et al. (2009) 

India 

64.2 6.8 106.1 

Forest 1880 - 1981 

Chhabra et al. (2003) 

10.9 0.041 

Total 47.5 Velayuthum et al. (2000) 

Total 329 24.0 73.1 Bhattacharyya et al. (2008) 

Indonesia Total 183 20.8 113.4 Shofiyati et al. (2010) 

Japan 

5.4 0.44 81.8 

Arable land 1979 - 1998 

Takata (2010) 

5.0 0.45 89.6 0.0001 

Forest 1950s - 1990s 24.3 2.18 90.0 Morisada et al. (2004) 

Total 29.3 2.63 89.9 Takata (2010) and Morisada et al. (2004) 

Table 10.1 Soil organic carbon change in selected countries in Asia 


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in Asia

China is a vast country – 6 percent of the global land area and 46 percent of the area of the region. Estimates 

in China of the totoal SOC storage in the 0 -1 00 cm soil profile show high variability, ranging from 50 to 183 Pg 

(Xie et al., 2007). The reasons for this variability include uncertainty about the area of croplands, the quantity 

of soil profiles and the methods applied for scaling up from soil profiles to a national level. Using data from 

the Second National Soil Survey carried out in the 1980s, Xie et al. (2014) estimated areas and SOC stocks as 

follows: paddy lands 29.87 million ha and 2.91 Pg C; uplands 125.89 million ha and 10.07 Pg C; forest 249.32 

milion ha and 34.23 Pg C; and grassland soils 278.51 million ha and 37.71 Pg C. 

Farmland and forests were found to have SOC sequestration rates of 23.62 and 11.72 Tg C yr –1, respectively, 

resulting in 0.472 and 0.234 Pg SOC respectively being accumulated in these soils during the period 1980 to 

2000. However, degradation of grassland depleted 3.56 Pg C over the same period. Thus during twenty years, 

a net amount of 2.86 Pg C was lost, approximately 3.4 percent of total SOC storage in China. 

With an artificial neural network model to link SOC change to six parameters - latitude, longitude, 

elevation, soil type, land use type, and original SOC in early 1980s - Yu et al. (2009) estimated an increase of 

260 Tg C occurred in Chinese cropland in the period 1980 to 2000 in the topsoils (0-20 cm). The increase of SOC 

content is mainly attributed to the large increase in crop yields and the increased residues retained in fields. 

By contrast, SOC storage in grassland is dwindling, especially in the northwest and southwest parts of China, 

mainly due to the degradation of grassland. Nationally, an area of 5.29 million ha grassland had degraded 1986 - 

1 

999 (Han and Gao, 2005). Unlike the grassland ecosystem, SOC storage in forestland has increased (see also 

above). On the forested area of 249.32 million ha, estimates of carbon sequestration range from 234 to 304 Tg 

SOC in the period of 1980 to 2000, mainly attributed to forest expansion and regrowth (Zhou et al., 2006; Xie 

et al., 2007). 

Other studies show conflicting results, with some showing recent Chinese soils as a net carbon sink. Tian 

et al. (2011) estimated SOC change using two process-based terrestrial ecosystem models with considering 

factors including climate change and land use change. They concluded that biomass and soils in China 

accumulated organic carbon at a rate of 0.121 and 0.094 Pg C yr -1 respectively from 1981 to 2000. Piao et al. 

(2009) adopted a bottom-up approach and showed consistent results (0.105 and 0.075 Pg C yr -1 , respectively) 

in the same period, while Houghton and Hackler (2003) reported net C loss from Chinese ecosystems (-0.008 

PgC yr -1 ) in 1990s, including loss in biomass (-0.028 Pg C yr -1 ) and net C sink in soils (0.029 Pg C yr -1 ). 

Velayutham and Bhattacharyya (2000) estimated soil organic C stock in different soil orders and different 

agro-ecological regions in India (Sehgal and Abrol, 1992). For the top 1 m depth, they estimated soil organic 

C stock as 47.5 PgC, which is double the previous estimates of Dadhwal and Nayak (1993) and Gupta and Rao 

(1994). The trend may, however, be negative as crop residues are widely used as fuel and fodder and not returned 

to the soil, which would result in a decrease in soil organic content. In Bangladesh, the average organic content 

is said to have declined by half, from 2 percent to 1 percent, over the past 20 years (Bangladesh, 1992). For the 

Indian State of Haryana, soil test reports over 15 years show a decrease in soil carbon (Chaudhary and Aneja, 

1991). Decreased organic content leads to: (i) degradation of soil physical properties, including water holding 

capacity, as has developed in India (Indian Council of Agricultural Research, personal communication); (ii) 

reduced nutrient retention capacity; and (iii) longer release of nutrients, including micronutrients, from 

mineralization of organic master. As a consequence of all these effects, there may be longer responses to 

fertilizer. 

The threat to soil organic carbon (SOC) change in Indonesia’s mineral and peat soils is mainly caused 

by deforestation, poor land management or intensive cropping, or by a combination of these factors 

(Hartanto et al., 2003; Lal, 2004). In lowland rice grown on mineral soils, SOC tends to be maintained or 

increased because of anoxic condition (Kyuma, 2004). Land use types greatly influence soil loss. Annual 

crop based systems where soil conservation measures are not practiced are associated with high erosion 


300 

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(Valentin et al., 2008). For peatland, the threat to SOC occurs when peat forest is cleared and drained (Page, 

Rieley and Banks, 2011). The change from saturated to unsaturated conditions leads to the enhancement 

of aerobic microbial activities. This is the main factor in the SOC loss which occurs through the resulting 

accelerated microbial decomposition (Hooijer et al., 2014; Agus and Subiksa, 2008). Peat fire is another cause 

of SOC loss. Maintaining a high water table is the key to reducing SOC losses through both these pathways. 

More details on the loss of soil carbon in Indonesia are included in the country case study section (10.5.2). 

Permanent meadows and pastures area occupy 73 percent of the land area of Mongolia, while forest and 

arable land area account for 7 percent and 0.4 percent, respectively (FAO, 2012). During the 1990s and 2000s, 

the forest area decreased (loss of 8.19×102 km 2 yr -1 ), and as a result forest biomass in Mongolia has decreased 

at a rate of 0.004 PgC yr−1 (Piao et al., 2009). Li et al. (2005) reported net ecosystem exchange in a Mongolian 

steppe under grazing using the eddy covariance technique and suggested that the steppe was almost carbon 

neutral. 

Forest occupies 69 percent of the land area of Japan, while arable land and permanent meadows and 

pastures area account for 12 percent and 2 percent, respectively (FAO, 2012). Total SOC stock in forest area 

was estimated separately by Morisada, Ono and Kanomata, (2004) and Ugawa et al. (2012) using a bottomup 

approach but with different data sets. Morisada et al. (2004) used 3 391 soil profile data sampled from the 

1950s to the 1970s and calculated the weighted average SOC stock as 90 Mg C ha -1 (0-30cm) and 188 Mg C ha -1 

(0 -1 00cm). Ugawa et al. (2012) compiled 4 km mesh (around 3 000 profiles) data sampled from 1999 to 2003 

and estimated average SOC stock as 69.4 Mg C ha -1 (0-30cm). Since the Japanese forest area has changed 

little during the last 40 years, and has even increased slightly, the difference of SOC values between the two 

studies could be attributed to the difference in the methodology of sampling. Assuming the average SOC 

stock represented the total forest area (25.0×106 ha), total SOC stock in forest area would be in the range of 1.7 

to 2.2 Pg C. Takata (2010) compiled soil survey data (1979 -1 998) to calculate SOC stock in Japanese arable soils, 

and reported that 0.44 Pg C in 1979 increased very slightly to 0.45 Pg C, mainly due to an increase of both area 

and stock per unit area in grassland. 

10.4.3 | Soil salinization and sodification 

As outlined above section (10.3.5), the threat of salinization/sodification in the region takes varying forms. 

In the semiarid and arid zones of central and west Asia, salt-affected soils are widely distributed. On the other 

hand, salt-affected soils are also developing in certain coastal areas in monsoon zones, caused mainly by salt 

water intrusion in South and Southeast Asia and by coastal tideland reclamation. Although the coastal area 

affected is relatively small, this could become a serious problem for lowland rice production. 

In the GLASOD study, the region is estimated to have 42 million ha affected by salinization, nearly all of 

which is located in the dry zone. Of this salinized area in the drylands, there are estimated to be approximately 

4 million ha in both India and Pakistan. Salinization is also a major problem on irrigated land: GLASOD 

estimates that 10 percent of irrigated lands in India are affected, 23 percent in Pakistan and 9 percent in Sri 

Lanka, although these percentages are probably overstated since some of the salinization results from saline 

intrusion into unirrigated land. 

GLASOD provides estimates of areas subject to strong salinization. These numbers are important, as by 

definition they refer to land abandoned and taken out of cultivation. However, there is sometimes a wide 

difference between GLASOD figures for strongly saline soils and country estimates, although it should be noted 

that some of these include naturally occurring saline soils. For India, country estimates range between 7 and 

26 million ha, all higher than the GLASOD value of 4 million ha. For Pakistan, there is better agreement; leaving 

aside three estimates of 9 -1 6 million ha, the GLASOD and six country estimates lie in the range 4-8 million ha. 

Two apparently independent surveys, by the Soil Survey of Pakistan and the Water and Power Development 

Authority, show relative agreement at 5.3 and 4.2 million ha, respectively. 


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in Asia

In Bangladesh, an extension inland of coastal soil salinity has been noted in recentyears. Lower river flows, 

reduced by upstream abstraction for irrigation, have proved insufficient to dilute and displace sea water. In Sri 

Lanka, small areas of light salinization have appeared on irrigated lands of the Mahaweli scheme; the problem 

has not yet reached serious proportions, but needs to be monitored. 

Estimates of the extent of saline soils need to be associated with the dates of survey. Through successful 

reclamation, the extent of saline soils has been reduced in some areas, particularly as a consequence of the 

series of Salinity Control and Reclamation Projects (SCARP) in Pakistan. For example in the Pakistan Punjab 

the area of waterlogged and saline soils, which had risen from 61 000 ha in 1960 to 68 000 in 1966, had been 

reduced to 23 000 ha by 1985 (Chopra, 1989). 

Tideland reclamation projects have been carried out for centuries on the western coast of Korea. Records 

show that tideland reclamation in Korea began at the Ganghwa island in Gyeonggi-Do in 1235 (the 22nd year of 

King Gojong, Goryeo Dynasty). Reclamation in the early years was on a small scale, but has expanded over the 

years. Large scale modern reclamation projects started in 1960s as part of the national development program. 

Some of these projects have been as substitutes for the more than 20 000 ha of farmland which have been 

converted each year into industrial estates or other urban purposes. Since 1945, 75 738 ha of tideland have been 

reclaimed for paddy fields in 185 project areas by the Korean government and private companies (Park, 2001). 

The main constraints to crop production on reclaimed tideland are soil salinity, a high water table with 

poor drainage, and an unfavorable soil chemical composition. Resalinization of the surface soil is caused by 

evapotranspiration during the dry season and by capillary rise of saline water from groundwater resources. 

High soil salinity in reclaimed tidelands needs to be managed by controlling the amount and quality of 

irrigation water (Jung, Joo and Yoon, 2002). Soil characteristics on reclaimed land change continuously as 

desalinization progresses. However, in general, newly reclaimed saline soils have poor chemical properties 

and weak physical soil characteristics of soil thickness, soil structure and water logging, so that it is difficult to 

grow crops economically (Park, 1991). 

10.4.4 | Nitrogen imbalance 

Balanced nutrient supply is essential for achieving high crop yields, but excessive or unbalanced nutrient 

inputs may pose risks to the environment, human health and ecosystems. Nutrient losses may occur via 

emission to the air or discharge to the water through runoff, leaching and erosion. 

Nutrient inputs in Asia vary considerably amongst countries, districts and farming systems. Nutrient inputs 

tend to be higher than in other regions, and the trend may continue. FAO has predicted that for the coming 

four decades 60 percent of the world population increase will be in Asia (FAO, 2014). Feeding this rapidly 

growing population while minimizing harm to the environment is of great importance for both Asia and the 

world. Innovations in policy, science and farming practice are urgently needed to achieve this goal. 

Nutrient cycles link agricultural systems to their societies and surroundings and create the need for decisions 

on tradeoffs. Inputs of nitrogen and other nutrients are essential for high crop yields, but downstream and 

downwind losses of these same nutrients diminish environmental quality and human well-being (Vitousek et 

al., 2009). In this section, the issue of nitrogen imbalances in agriculture in Asia is highlighted as an important 

threat to soil. 

One study found that nutrient balances differ among Asian countries, varying generally with the level of 

economic development (Vitousek et al., 2009). The study examined six countries representing developing 

(China and India), developed (Japan and South Korea) and least developed countries (Laos and North Korea). 

China had the highest input of nitrogen (505 kg N ha -1 of arable land), which is nearly ten times levels applied 


302 

in Asia

in Laos (59 kg N ha -1 ). On the other hand, the mean N output was 108 kg N ha -1 in China, while North Korea 

achieved half that level (44 kg N ha -1 ). The N input was dominated by fertilizer application, especially in the 

booming economies of China and India, where the fertilizer N inputs accounted for 76 percent and 62 percent 

respectively of the total N inputs. In the least developed countries, such as Laos and North Korea, the N input 

mainly consists of animal manure, crop straw and biological N fixation. For developed countries, such as Japan 

and South Korea, the N inputs were more evenly contributed by fertilizer and manure. However, because of 

relatively high N input, higher N losses were observed in China, Japan and South Korea, compared with the 

least developed countries. The nitrogen use efficiency (NUE) was relatively high in Laos and North Korea, at 

the expense of depleting the soil N. The average soil N depletion rate ranged between 10–30 kg N ha -1 yr -1 in 

Laos and North Korea, while in China soil N accumulated by 26 kg N ha -1 yr -1 . 

In the least developed countries, manures were considered as precious nutrient resource for crop 

production. This was also the case in China before the 1980s, when artificial fertilizer was not subsidized. 

However, from the 1980s on, fertilizer was more widely used than manure in China. The amount of fertilizer 

applied in China far exceeds rates in other countries in Asia, and even exceeds rates in the United States and 

EU (Ju et al., 2009; Vitousek et al., 2009). In China in 2005 lack of regulation led to more than half of manure 

being discharged untreated to water bodies (Ma et al., 2010). By contrast, the environment is highly protected 

in Japan. More than 80 percent of manure is treated before being applied to crop land (Mishima, 2012). The 

N inputs and outputs also showed large variations between different crops. The highest nutrient inputs and 

accumulation in cash crop production were observed in China (Yan et al., 2013). As both livestock and cash crop 

production are expected to increase considerably in Asia in the future as demand rises from a fast growing and 

increasingly urbanized population (FAO, 2014), N use is likely to increase but there may be a more intelligent 

and sustainable approach to N inputs and outputs. 

Nitrogen use efficiency (NUE) is defined by the N output as crop production divided by the N inputs that are 

added through fertilizer, biological N fixation and N deposition (Ma et al., 2010). In a study across Asia and the 

Middle East, national average NUE ranged from 131 percent in Laos to 19 percent in the United Arab Emirates. 

The NUE was more than 100 percent in Laos, Nepal and Myanmar, because these countries relied to 

a great extent on recycled N inputs such as manure and crop straw. The NUE showed a reverse trend of N 

surplus among countries, decreasing from high N depletion countries such as Nepal and Laos to the high N 

surplus countries such as the United Arab Emirates and China. None of the crop production systems were 

sustainable in high N depletion or surplus countries. Although lower N losses and higher NUE were observed 

in the least developed countries, the continuous N depletion will limit crop yields and food production. In N 

surplus countries, the high levels of N accumulation may lead to higher N losses and consequently to serious 

environmental problems (Guo et al., 2010; Liu et al., 2013). 

The average N losses varied among countries, from 23 kg N ha -1 yr -1 in Afghanistan to 327 kg N ha -1 yr -1 in China 

(Figure 10.3). This was due to the differences in N input rate, the crop production structure and the area of 

arable land. For a variety of reasons, large areas of arable land are not presently cultivated in Afghanistan, Iraq, 

Mongolia and Syria. The contribution of the different N loss pathways varied among countries, for example, 

ammonia emission contributed to 50 percent of the N losses in Philippines, but to only 22 percent of the N 

losses in Laos. 

In conclusion, N inputs and outputs vary considerably amongst countries in Asia, with variations mainly 

attributable to levels of development and to policies. The nutrient imbalance in Asia could have a large impact 

on crop production and on the environment. These impacts may increase in the future with the further rapid 

development of livestock and cash crop production. Both the N imbalance and the N losses can be improved 

greatly without any sacrifice of crop yield. For example, balanced N fertilizer application and maximum 

manure N application standards have already proved effective in the EU (Velthof et al., 2009). Recent studies 


303 

in Asia

show that crop yields in China can be increased by 20-30 percent with no increase in input of fertilizer simply 

by using ISSM technologies (Chen et al., 2014). However, in the least developed countries increased use of 

fertilizer would greatly boost crop yields, as has happened in China during the past 30 years. For the developed 

countries, promoting mixed crop and livestock production systems could be the best choice to mitigate 

nutrient losses 

Figure 10.3 Nitrogen surplus or depletion, and nutrient use efficiency in crop production in Asia and the Middle East in 2010. 


10.5.1 | Case study for India 

The area of India is 328.2 million ha, of which 141 million ha is under cultivation. The country is bordered by 

the Arabian Sea in the west, the Bay of Bengal in the east, and the Himalayas in the north. Physiographically, 

India is divided into four broad divisions: (i) Himalayan Range (Northern Mountains); (ii) Hill regions, Indian 

Peninsula and Eastern Plateau; (iii) the great Indo-Gangetic Plains and Coastal Plains; and (iv) the Islands 

(Singh, 1971). India is endowed with diverse climates and there are three distinct main seasons (rainy, winter 

and summer). The country is influenced by monsoonal type climate and rainfall. Annual rainfall varies from 

less than 100 mm in the cold desert area of Jammu & Kashmir, Lahaul Spiti and the Thar Desert of Rajasthan 

to over 11 000 mm in Cherrapunji in Meghalaya. Mean annual temperatures in the country vary from 8 to 

28°C. Variation in the mean summer and mean winter temperature in the northern region is

Sub-Regions. Inceptisols are the dominant soils covering 39.75 percent of the total area, followed by Entisols 

(28.08 percent), Alfisols (13.55 percent), Vertisols (8.52 percent), Aridisols (4.28 percent), Ultisols (2.51 percent), 

Mollisols (0.4 percent) and others (2.92 percent) (Bhattacharyya et al., 2013). 

Degraded and wastelands of India 

Velayutham and Bhattacharyya (2000) reported that the total area subject to soil degradation in India is 45.9 

percent. Of this, 37.0 percent is affected by water erosion, followed by wind erosion (4.0 percent), salinization 

(2.2 percent), loss of nutrients (1.1 percent) and waterlogging (1.6 percent). Land not fit for agriculture (icecaps, 

salt-flats, arid mountain and rock outcrops) constitutes 5.5 percent of the total area. About 27.5 percent 

of soils have no degradation problem and the remaining 9.8 percent of the area is classed as ‘stable terrain’, 

e.g. natural conditions. Recently the National Academy of Agricultural Sciences (NAAS) inventoried the soil 

degradation status of the country based on reconciliation of databases gathered from different organizations 

(Figure 10.4 and Table 10.2). This confirmed that soil erosion by water is the most serious threat, affecting 82.57 

percent of the total area, followed by wind erosion (12.40 percent), acidic soils (17.94 percent), salt-affected 

soils (6.74 percent), waterlogged soils (0.88 percent) and mining and industrial waste (0.19 percent). 

Figure 10.4 Degradation and wastelands map of India. Source: ICAR and NAAS, 2010. 


305 

in Asia

Open forest 

Degradation type 

Arable land (million ha) 

(10 tonnes ha -1 yr -1 ) 

73.27 9.30 

Soil Loss Map of 

India–CSWCR&TI 

Wind erosion (Aeolian) 12.40 – 

Wind Erosion Map of 

India–CAZRI 

Chemical degradation 

Sub-total 85.67 9.30 

Exclusively 

salt- affected soils # 

Salt-affected and 

water eroded soils 

5.44 

1.20 0.10 

Salt-Affected Soils 

Map of India, CSSRI, 

NBSS&LUP, NRSA and 

others 

Exclusively acidic soils 

5.09 - 

(pH< 5.5) # 

Acid Soil Map of India 

NBSS&LUP Acidic 

(pH < 5.5) and water 

Acidic soils (pH< 5.5) 

and eroded soils 

5.72 7.13 

Physical degradation 

Sub-total 17.45 7.23 

Mining and industrial 

waste 

0.19 

Wasteland Map of 

India–NRSA 

Waterlogging 

(permanent surface 

0.88 

inundation) $ 

Sub total 1.07 

Total 104.19 16.53 

Grand total 

(arable land and open forest) 

120.72 

Table 10.2 Harmonized area statistics of degraded and wastelands of India. Source: ICAR and NAAS, 2010. 

Notes: FSI (1999) was used to exclude degraded land under dense forest; Unculturable Wastelands: Barren rocky/stony waste: 6 M ha, are the source for 

runoff water and building material; Snow covered/Ice-caps: 6 M ha, are best source of water and are not treated as wastelands 

#For acid soils, areas under paddy growing and plantation crops were also included in the total acid soils 

$ 

Sub-surface waterlogging not considered. 


306 

in Asia

10.5.2 | Case study for Indonesia 

Topography, climate, soil parent materials and anthropogenic factors determine the various types and 

degree of soil threats in Indonesia. These threats include soil carbon depletion both for mineral and organic 

soils, erosion by water, soil contamination, soil acidification and nutrient imbalance. Soil organic carbon 

depletion in peat is the most significant threat. 

Soil erosion threats are found in almost all hilly and mountainous landscapes of the Indonesian archipelago 

due to the very high (>2000 mm) annual rainfall over most of the area (83 percent). Most other sloping areas 

with lower rainfall are also affected by erosion due to the high intensity of the monsoonal rainfall during the 

rainy season (Agus, Amien and Sutono, 2002). Besides causing erosion, the high rainfall also leads to leaching 

of basic cations and hence to soil acidification. The main problems associated with managing acid soil are low 

pH, P-fixation, low basic cation concentration, low cation exchange capacity (CEC) and toxicity of soluble Al 

and Fe. 

The total area of Indonesian acid upland soils (pH

palm fronds increased SOC, especially in the 20 percent of the plantation area receiving an equivalent of 4.8 

Mg C ha -1 yr -1 from palm fronds (Haron et al., 1998). If degraded land is converted to plantation, it was found 

to gain about 30 percent SOC stock (Murty et al., 2002; Germer and Sauerborn, 2008). In East Kalimantan, 

natural regeneration from Imperata grassland to secondary forest increased soil carbon content by 14 percent, 

from 14.5 g kg -1 to 16.5 g kg -1 (van der Kamp, Yassir and Buurman, 2009). In Java, agricultural top soil with 

continuous rice cropping accumulated more than 1.7 Tg C per year over the period of 1990-2000 (Minasny 

et al., 2012). Kyuma (2004) also suggested that SOC in newly established paddy rice soils tends to increase 

with time. Strategies to increase the soil carbon pool include soil restoration and woodland regeneration, notill 

farming, cover crops, nutrient management, manuring and sludge application, improved grazing, water 

conservation and harvesting, and efficient irrigation (Lal, 2003). 

The loss of carbon from organic soils 

Indonesian peatland is estimated to cover about 14.9 million ha (Ritung et al., 2011). It is a massive store 

of carbon, storing around 27 Pg C (Agus et al., 2013). This huge C store is formed under saturated condition 

in concave areas with the annual rate of peat formation in the range of 0-3 mm thickness (Noor, 2001) or 

equivalent to zero to about 1.5 Mg C ha yr -1 carbon accumulation (Agus and Subiksa, 2008; Parish et al., 2007). 

This relatively slow process has led to the formation of peat domes that began between 6 800 and 4 200 years 

ago (Andriesse, 1994). Some formations may be as old as 26 000 years (Page et al., 2002). 

Carbon stock in peat ranges from 420 to 820 Mg C ha -1 . Peat depths range from 0.5m to over 10 m (Agus, 

Hairiah and Mulyani, 2011) and the peat C stock is strongly determined by peat thickness (Warren et al., 2012). 

Under saturated natural conditions, peat C is slowly emitted in the forms of CO 2 

and CH 4 

due to anaerobic 

microbial activities. As peat forest is cleared and drained, the peats become unsaturated and CO 2 

emission 

escalates at a pace far exceeding the rate of sequestration. Currently, from the 14.9 million ha classified as 

peatland, forested areas amount to just over half (52 percent), shrub cover is about 21.7 percent, and the rest 

is under agriculture or settlements. Peat soil C loss runs in parallel to the level of local development. Peatland 

areas of Sumatra and Kalimantan have been drained on a wide scale, and hence are fast losing carbon. On the 

other hand, Papua peatlands are mostly conserved (Figure 10.5). 

The major processes of carbon loss from drained peatland are (i) peat decomposition under aerobic 

condition and (ii) peat fire. Natural phenomena such as lengthy droughts aggravate peat fire risks (Page et al., 

2002; IPCC, 2014; Parish et al., 2007; Agus and Subiksa, 2008; Wosten et al., 2008). 

The literature varies in its estimates of peat carbon loss through decomposition. One source (IPCC, 2014), 

based on research data from Malaysia and Indonesia, presented peat CO 2 

-C emission factors based on land 

cover classes. In this study, degraded forest is expected to emit as much as 5.3 Mg CO 2 

-C ha -1 yr -1 , while various 

agricultural and forestry uses emit a wide range of CO 2 

-C as shown in Table 10.3. The 95 percent confidence 

interval was very wide for each land cover class, which suggests the need for site specific emission factors. A 

case study in Riau Province (Husnain et al., 2014) showed insignificant differences in peat emissions under an 

oil palm plantation, an Acacia plantation, a secondary forest and a rubber plantation. The emission rates were 

18.0 ± 6.8; 16.1 ± 5.3; 16.6 ± 6.8; 14.2 ± 4.6 Mg CO 2 

-C ha -1 yr -1 , respectively. For bare land sites, the rates measured 

lay between 15.2 ± 8.2 and 18.2 ± 6.5 Mg CO 2 

-C ha -1 yr -1 . The findings of other studies (Marwanto and Agus, 2014; 

Dariah, Marwanto and Agus, 2014) were similar to those of the IPCC (2014), with emissions from oil palm 

plantations on peat soils of about 10 to 11 Mg CO 2 

-C ha -1 yr -1 . 

A study in Central Kalimantan Province revealed a linear relationship between average water table depth 

and peat CO 2 

emission. Degraded drained forest was found to emit about 2.7 Mg CO 2 

-C ha -1 yr -1 in areas with 

zero mean water table depth, and about 15 Mg CO 2 

-C ha -1 yr -1 for areas where average water table depth 

was 1 m (Hooijer et al., 2014). Regardless of the variations found amongst the studies, the values recorded 

demonstrate rapid carbon depletion in drained peatland. 


308 

in Asia

Peat fire is another important process that may cause a huge amount of carbon loss in a short period of 

time, and the literature is in agreement on the importance of peat fire as one of the main causes of peat 

carbon loss (Page et al., 2002; Hooijer et al., 2014). However, determination of peat C loss through peat fire is 

a future research challenge, especially with regards to gathering firm data, for example on the volume (area 

and depth) of burnt scars (IPCC, 2014). 

Land cover type 1 

Drained forest land 

and cleared forest land 

(shrubland) 

Plantations, drained, 

unknown or long 

rotations 


short rotations, 

e.g. acacia 


oil palm 

Plantations, shallowdrained 

(typically less 

than 0.3 m), typically 

used for agriculture, 

e.g. sago palm 

Cropland and fallow, 

drained 

Cropland, drained, 

paddy rice 

Emission 

(Mg CO 2 

-C ha -1 yr -1 ) 

95 percent Confidence level 2 

(Mg CO 2 

-C ha -1 yr -1 ) 

5.3 -0.7 9.5 

15 10 21 

20 16 24 

11 5.6 17 

1.5 -2.3 5.4 

14 6.6 26 

9.4 -0.2 20 

Grassland, drained 9.6 4.5 17 

Table 10.3 Emission factors of drained tropical peatland under different land uses and the 95 percent confidential interval. Source: 

IPCC, 2014. 

1) Emission of primary peat forest is assumed to be zero. 

2) Some confidence intervals contain negative values because calculation was based on error propagation of uncertainties. 

However, all underlying CO 2 

fluxes were positive. 


309 

in Asia

Figure 10.5 Indonesian peatland map overlaid with land cover map as of 2011. Source: Wahyunto et al., 2014. 

10.5.3 | Case study for Japan 

The islands which make up Japan are located in the one of the most active parts of the Circum-Pacific Ring 

of Fire. The major islands are Hokkaido, Honshu, Shikoku and Kyushu. There are 110 active volcanoes in Japan 

(16 volcanoes in Hokkaido, 38 volcanoes in Honshu and 11 volcanoes in Kyushu). Large quantities of tephras 

from volcanoes have been deposited on the Pleistocene terraces, constituting a main parent material of 

Japanese soil (Andosols). Japan’s total land area is about 378 000 km 2 . About 72 percent of Japan’s land area is 

mountainous, and rivers are characterized by their steep gradients and relatively short lengths. About twothirds 

of the total land area consists of forest. The plains cover only about 28 percent of the total land area. 

Most plains are located along the seacoast. Arable lands account for 12.1 percent of the total land surface, 

mainly distributed in the plains. The climate of Japan is influenced by a monsoonal flow that carries moist air 

from the Indian Ocean and Pacific Ocean. In general, Japan has four distinct seasons: spring (March to May), 

summer (June to August), autumn (September to November) and winter (December to February). 

Japan’s arable land covers 4.5 million ha, with paddy fields accounting for just over half the area (54 percent). 

Grey lowland soils (Fluvic Hydragric Anthrosols or Gleyic Fluvisols (FAO/IUSS/ISRIC, 2006) comprise the largest 

cultivated soil area; followed by Gley soils (Gleyic Fluvisols), Andosols (Aluandic or Silandic Andosols), Brown 

forest soils (Haplic Cambisols), Brown lowland soils (Haplic Fluvisols), and Wet Andosols (Gleyic Andosols). 

Urban sprawl and other changes in land use (including abandonment of cultivation) led to a shrinking of the 

agricultural land area by about 1 million ha between 1973 and 2001. Urbanization and consequent soil sealing 

advanced into flat lowland areas, largely into paddy fields on Grey Lowlands soils and Gley soils. In addition, 

loss of Andosols to expanding urbanization was widely observed over the flat upland fields in the middle part of 

Hunshu islands. By contrast, upland fields on steep slopes in the western part of Japan, largely on distributed 

Brown Forest Soils, were simply abandoned (Takata et al., 2011b). 

Soil organic carbon change 

Spatio-temporal variations in soil organic carbon (SOC) content in arable land were evaluated by both 

model-based (Yagasaki and Shirato, 2014) and monitoring-based (Takata, 2010) approaches. In the modelbased 

approach, SOC stock change was simulated using the original Rothamsted Carbon model (Coleman 

and Jenkinson, 1996) and two modified Rothamsted Carbon models (Shirato, Yagasaki and Nishida, 2011; 

Takata et al., 2011a). The rate of change in the total SOC stock in Japanese agricultural lands evaluated with 

10year intervals was estimated to be -0.95 Tg C yr -1 between 1980 and 1990. A greater loss of SOC, equal to -1.06 

Tg C yr−1, was found subsequently for the period from 1990 to 2000. 


310 

in Asia

An agricultural soil monitoring project named ‘Basic Soil-Environmental Monitoring in Japan’ has been 

conducted since 1979. This soil monitoring project has taken readings at repeated five year intervals at about 

20 000 fixed points. The rate of change in the total SOC stock evaluated by this monitoring was 2.3 Tg C yr - 

1 

from 1979 to 1989. The project found that SOC increased in arable land. However, the monitoring detected a 

greater loss of SOC for a later period (1989-1998), equal to -2.3 Tg C yr−1. During this period, the agricultural 

land area decreased from 5.4 to 5.0 million ha. At the same time, soil carbon content in arable land gradually 

rose from 88 to 90 tonnes C ha -1 . These results indicated that the decline of SOC stock in Japanese arable land 

1989-1998 was mainly influenced by the fluctuation in the arable land area. 

Spatial variation of SOC content in forest soils has been monitored since 2005. The mean SOC content of 

forest soil at 0-30cm was 69.4 tonnes C ha -1 (Ugawa et al., 2012), lower than the value of 90 tonnes C ha -1 in 

arable land. 

Heavy metal contamination 

Rapid industrialization in Japan during the 1960s polluted arable soil with heavy metals such as cadmium 

(Cd), Copper (Cu), and Arsenic (As). There were four main pollution sources: mining activity, factories 

and incinerators, fertilizer, and precipitation and irrigation water (Makino, 2010). In 1970, the Japanese 

government enacted the Agricultural Land Soil Pollution Prevention Law to regulate heavy metal pollution. 

The law designated Cd, As, and Cu as hazardous substances to be regulated. The allowable limitation of Cd 

was set in terms of the Cd concentration in rice grains (1 mg kg -1 ). The allowable limitations on As and Cu were 

set to 15 (1M HCl soluble) and 125 (0.1M HCl soluble) mg kg -1 soil, respectively. The amount of bioavailable Cd 

in soil is affected by many factors, so setting an allowable concentration in terms of the soil Cd content is 

impractical (Asami, 1981). The area of polluted arable land was assessed as 7 592 ha (Cd, 7 050 ha; Cu, 1 405 ha; 

As, 391 ha). About 92 percent of the total polluted area has been remediated by uncontaminated soil dressing 

(MOE, 2014). 

Radioactive Cs contamination 

As a result of the accident at the Fukushima Dai-ichi Nuclear Power Station (FDNPS) operated by the 

Tokyo Electric Power Company, radioactive cesium (Cs) was released into the surrounding environment. To 

determine the extent of decontamination required in arable land and to consider management options, the 

Ministry of Agriculture, Forestry and Fishery (MAFF) surveyed and measured soil Cs concentrations in 3 461 

agricultural fields, and used these data to construct a distribution map of radioactive Cs concentration in 

agricultural soil in eastern Japan (Takata et al., 2014). The distribution map of radioactive Cs concentration in 

agricultural soil is shown in Figure 10.6. 

Contamination level 4 (>25 000 Bq kg -1 ) was observed only in a 20 km evacuation area (EA) surrounding 

FDNPS (EA-20km) and the Deliberate Evacuation Area (DEA); 77 percent of the level 4 areas were observed 

in the EA-20km. Farmers of level 4 contaminated fields were advised to solidify their topsoil with a fixation 


311 

in Asia

Fukushima 

Legend 

(Bq kg -1 ) 

0 - 500 

500 - 1000 

1000 - 5000 

5000 - 10000 

10000 - 25000 

25000 - 50000 

50000< 

Figure 10.6 Distribution map of radioactive Cs concentration in soil in Fukushima prefecture (reference date of 5 November, 2011). 

Source: Takata et al., 2014. 

agent to prevent scattering of contaminated soil during a topsoil removal operation. Contamination level 3 

(10 000–25 000 Bq kg -1 ), which indicated a need to remove topsoil, was distributed in only the EA-20km and 

DEA, with more than half (55 percent) in the EA-20km. Eighty percent of fields with contamination level 2 

(5 000–10 000 Bq kg-1) were distributed in the evacuation zone, with the remaining 20 percent distributed 

in the non-evacuation zone in Fukushima Prefecture. Paddy fields with contamination level 2 (2 100 ha) have 

three options for decontamination: topsoil removal, fine-textured topsoil removal using water, and topsoil 

burying. Upland fields, orchards, and meadows that are at contamination level 2 (1 200 ha) have two options 

for decontamination: topsoil removal and topsoil burying. 

Nutrient imbalance 

The soil surface nitrogen (N) and phosphate (P) balance in Japanese arable land has been improving 

(Mishima, Endo and Kohyama, 2010a; 2010b). These values serve as the index of the impact of arable land 

on the environment and of the farm-gate balance of nutrients. The soil surface N (and P) balance is defined 

as the total N (P) input (N kg ha -1 ) minus the total N (P) output (N kg ha -1 ). Chemical fertilizer application in 

Japan declined continuously during the period from 1985 to 2005. The application rate of livestock manure 

also peaked in 1990 and declined thereafter. Crop production, however, remained constant during this period. 

Between 1985 and 2005, the surplus N and P (positive value of soil surface N balance) declined from 89.9 to 49.3 

kg N ha -1 and from 153 to 105 kg P ha -1 respectively (Mishima, Endo and Kohyama, 2010a; 2010b). However, this 

trend was not consistent at the regional level because organic amendment applications were largely related 

to the availability and movement of livestock excreta (Mishima, Endo and Kohyama, 2010a) and to soil type 

(Leon et al., 2012). High surplus P and low crop P uptake compared with N, P input for crop production could be 

reduced. This limited negative environmental effects such as eutrophication of soil and water and conserved 

limited P resources. 


312 

in Asia

The Basic Soil-Environmental Monitoring Project found excess soil Ca in paddy fields, upland fields and 

orchards during 1979-1998 (MAFF, 2008), but also a gradual increase in soil Mg deficit over the same period. 

Thus the balance of Ca and Mg has been deteriorating in Japanese arable land. In addition, the soil pH of paddy 

fields gradually decreased from 5.8 to 5.7 1979-1998 (MAFF, 2008). 


Spatial estimation of soil loss from arable land at national scale was carried out using the Universal Soil 

Loss Equation (USLE) and environmental inventories (Kohyama et al., 2012). Hourly rainfall and runoff factors 

(R: resolution 1 km) were calculated by the amount of precipitation analysed by radar-AMeDAS. Topographic 

factors (LS) are shown in Figure 10.7, calculated using a digital elevation model (resolution 50 m) and ALOS 

satellite imagery. The soil erodibility factor (K) of arable land was calculated using the physico-chemical 

soil properties (soil texture, soil organic matter content, etc.) as measured in the Basic Soil-Environmental 

Monitoring Project (Taniyama, 2003). The K factor of arable land was determined by soil series groups. The 

K factor was relatively higher in clayey lowland soil group than in humic Andosol groups. The cover and 

management factor (C) of each crop was determined by Taniyama (2003), and was delineated using the agroenvironmental 

census data map (Kohyama et al., 2003). 

0-‐4 

5-‐9 

10-‐14 

15-‐19 

20-‐24 

25-‐29 

30-‐34 

35-‐40 

40-‐ 

0-‐20 

20-‐40 

40-‐60 

60-‐ 

Topographic factor (LS) Soil erodibility factor (LS; kg h MJ -‐1 mm -‐1 ) 

0-‐0.1 

0.1-‐0.2 

0.2-‐0.3 

0.3-‐ 

I 

II 

III 

IV 

V 

VI 

Non-‐arable land 

Cover and management factor (C) 

ClassificaBon of esBmated soil loss 

venerdì 20 novembre 15 

Figure 10.7 Distribution map of the parameters of USLE and classification of estimated soil loss. Class I: less than 1 tonnes ha -1 yr -1 ; 

Class II: 1-5 tonnes ha -1 yr -1 ; Class III: 5-10 tonnes ha -1 yr -1 ; Class IV: 10-30 tonnes ha -1 yr -1 ; Class V: 30-50 tonnes ha -1 yr -1 ; Class VI: more 

than 50 tonnes ha -1 yr -1 . Source: Kohyama et al., 2012. 


313 

in Asia

Loss of soil in Japanese arable land was categorized into six classes: Class I; less than 1 tonnes ha -1 yr -1 , Class 

II; 1-5 tonnes ha -1 yr -1 , Class III; 5-10 tonnes ha -1 yr -1 , Class IV; 10-30 tonnes ha -1 yr -1 , Class V; 30-50 tonnes ha -1 yr - 

1 

, Class VI; more than 50 tonnes ha -1 yr -1 . The proportion of soils in these classes was: 43 percent in Class I; 18 

percent in Class II; 9 percent in Class III; 12 percent in Class IV; 5 percent in Class V; and 14 percent in Class VI. 

Highly erodible zones were mainly distributed in areas of western Japan which are characterized by complex 

topography and heavy precipitation. 

10.5.4 | Case study of greenhouse gas emissions from paddy fields 

Rice is the staple crop for the majority of the world’s population. In Asia, rice cultivation areas roughly 

account for 89 percent of the global total (Yan, Akimoto and Ohara, 2003). While rice production is thus 

vital for feeding the world’s population, it is also an important source of greenhouse gas emissions, notably 

methane (CH 4 

) and nitrous oxide (N 2 

O). CH 4 

is converted from substrate by methanogenic bacteria in strictly 

anaerobic environments, while N 2 

O is an intermediate production of nitrification and denitrification. Both 

of these two gases possess considerably greater infrared absorbing capability than carbon dioxide (CO 2 

) on a 

mass basis: 25 times for CH 4 

and 298 times for N 2 

O. 

Using the tier 1 method described in the 2006 Intergovernmental Panel on Climate Change (IPCC) Guidelines 

for National Greenhouse Gas Inventories (IPCC, 2006), and country-specific estimates of rice harvest area and 

data on agricultural activities, Yan and colleagues estimated that global CH 4 

emission for 2000 was 25.6 Tg 

CH 4 

yr -1 , with a 95 percent uncertainty range of 14.8 to 41.7 Tg CH 4 

yr -1 , considerably lower than earlier estimates 

(Yan et al., 2009). Rice paddies in monsoon Asia countries contributed far and away the largest share of these 

emissions, estimated at 23.7 Tg CH 4 

yr -1 . China, with an amount of 7.41 Tg CH 4 

yr -1 , was estimated to be the 

largest CH 4 

emission country, followed by India, Bangladesh, Indonesia, Vietnam, Myanmar and Thailand. The 

areas with the greatest emission intensity were the delta regions of large rivers in Bangladesh, Myanmar and 

Vietnam, the island of Java in Indonesia, central Thailand, southern China and the southwestern portion of 

the Korean peninsula (Figure 10.8). 

CH 4 

emission from rice fields is the net result of three processes: production, oxidation and transport. 

Three main factors affect one or more of these processes: organic amendment, the water regime during 

the rice-growing season, and water status in pre-season. Appropriate agricultural management of organic 

amendments and the water regime should therefore be promoted to mitigate CH 4 

emissions. Techniques 

could include off-season straw incorporation and midseason drainage (Yan et al., 2009). 

Unlike CH 4 

, N 2 

O emission from rice paddies has been found to be much lower than from upland crops, 

because under a flooded environment, nitrification is weak and denitrification proceeds to the end step 

with N 2 

as the dominant product. The majority of N 2 

O emissions from rice paddies usually occur shortly after 

the mid-season drainage and nitrogen fertilizer additions (Yan et al., 2000). The availability of substrates 

and soil moisture condition are critical controlling factors for N 2 

O emission since they affect the activity of 

nitrifiers and denitrifiers. The fertilizer-induced N 2 

O emission factor for rice fields averages approximately 0.3 

percent, probably fluctuating dependent on the water regime (Akiyama, Yagi and Yan, 2005). Mitigation of 

N 2 

O emissions from paddy fields should also not be overlooked because of the excessive nitrogen fertilizer 

consumption by rice production. Since a tradeoff relationship between CH 4 

and N 2 

O emissions from rice 

paddies is frequently observed, any mitigation options pursued should take their comprehensive global 

warming potential fully into account. 


314 

in Asia

Figure 10.8 Estimate CH 4 

emission from rice paddy in Asia. Source: Yan et al., 2009. 

10.6 | Conclusion 

In Asia, management of land and water resources has been identified as one of the priority ways to 

achieve sustainable food security by raising land productivity, reversing land degradation and water loss, and 

increasing biodiversity and the quality of the environment. Asian countries have also committed themselves 

to strengthening regional cooperation and national capacities to develop a more integrated approach to the 

management of natural resources. An integrated approach is needed to improve the ability of countries to 

plan and monitor the better use and management of their land resources to increase agricultural productivity 

while maintaining land and environmental quality. 

However, since the GLASOD and ASSOD projects of the 1980s and 1990s, no extended assessment of the 

status of soil resources has been carried out in the region. There have been extensive scientific communications 

amongst experts in the region, including the activities of the East and Southeast Asia Federation of Soil Science 

Societies (ESAFS). Based on the above finding, a provisional assessment is made of the status and trend of the 

10 soil threats in order of importance for the region. At the same time an indication is given of the reliability of 

these estimates (Table 10.4). 


315 

in Asia

However, a number of the country reports that contributed to this chapter emphasized that rapid socioeconomic 

change and resulting changes in land use and its management, as well as climate change, have 

had great impacts on the soil resources in Asian countries. Therefore, there is a need to conduct a new and 

extensive assessment of changes in soil resources in the region. 

Responding to this need, a regional conference on soil information was held in Nanjing, China, on 

February 2012 to share the latest soil information and knowledge about advanced science and technologies 

on soil resources in Asian region. The conference recognized the benefits to be gained from further sharing 

of information and data on soil surveys, soil mapping and capacity development. The conference saw the 

establishment of the initial Asia Soil Partnership (ASP) and the signing of the Nanjing Communiqué (GSP, 

2012), which put the following goals as the priorities: 

1. sharing and transferring soil knowledge and new technology within and beyond the region 

2. providing soil information to all those with an interest in the sustainable use of soil and land resources 

3. building consistent and updated Asian soil information systems and starting to contribute to the global 

soil information system through initiatives such as GlobalSoilMap.net 

4. training new generations of experts in soil science and land management 

After the endorsement of the global Plans of Action in GSP, the next step in the Asian Region is the 

development of the regional implementation plan for sustainable soil management, which can translate the 

plans in the Nanjing Communiqué into practice. 


316 

in Asia

Threat 

to soil 

function 


Organic 

carbon 

change 

Salinisation 

and 

sodification 

Nutrient 

imbalance 

Contamination 

Summary 

Serious water erosion occurs 

in regions with dry and wet 

seasons covering South Asia 

to East Asia, particularly in 

the hilly and mountainous 

landscapes. However, it is 

of little concern for wellestablished 

forests and 

paddy fields. 

Wind erosion is 

concentrated mainly in the 

most western and northern 

arid and semi-arid regions 

of Afghanistan, Pakistan, 

India, and China. 

Increase in crop yield retains 

soil organic carbon (SOC) 

in croplands of East and 

Southeast Asia. Whereas, 

SOC is decreasing in South 

Asia, because crop residues 

are widely used as fuel and 

fodder, and not returned to 

the soil. 

The degradation of 

grassland has caused great 

losses of SOC stock. 

The threat of salinisation/ 

sodification in the Asia 

region is widespread but 

variable. In semiarid and 

arid zones of central Asia, 

salt-affected soils are widely 

distributed. On the other 

hand, salt- affected soils 

are developed in certain 

coastal areas in monsoon 

zones, mainly by salt water 

intrusion in South and 

Southeast Asia. 

Negative soil nutrients 

balances have been 

reported for N, P, K and 

micronutrients in many…” 

South Asian countries. 

Whereas, large excess of 

nutrients, in particular 

N, causes serious 

environmental problems in 

other countries. 

Rapid urbanization, 

industrialization, and 

intensive farming causes 

contamination of heavy 

metals (Cd, Ni, As, Pb, 

Zn, etc.) and pesticides in 

various parts of Asia, which, 

in turn, poses a serious risk 

to human health. 


Confidence 



317 

in Asia


and land 

take 


acidification 

Compaction 

Waterlogging 

Loss of soil 

biodiversity 

Rapid urbanization and 

development of mega-cites 

significantly increased the 

rate of impervious surface 

area (ISA). Asia region has 

the largest ISA within the 

global regions. 

There is substantial area 

of acid soils distributed in 

tropical and subtropical 

regions of Asia, mainly in 

Southeast Asia, parts of East 

and South Asia. 

This is mainly caused by 

unbalanced and unsuitable 

application of chemical 

fertilizers. Distribution 

of acid sulphate soils in 

tropical Asia also limits crop 

production. 

Mechanization of land 

management has increased 

compaction of surface 

soil and/or subsoil in 

cropland, grassland and 

timber forests. Increase in 

livestock trampling is also a 

major cause of surface soil 

compaction in grassland and 

hilly region. 

Anthropogenic activates 

such as poor drainage 

system and deforestation in 

the upstream areas increase 

the threat to waterlogging 

in the flood prone areas. 

Limited information is 

available for soil biodiversity 

in Asia. Some reports show 

high microbial biodiversity in 

the soils of organic farming 

lands. 

Table 10.4 Summary of Soil Threats Status, trends and uncertainties in Asia 


318 

in Asia

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in Asia

11 | Regional assessment of soil 

changes in Europe and Eurasia 

Regional Coordinator/Lead Author: Pavel Krasilnikov (ITPS/Russia) 

Contributing Authors: Irina Alyabina (Russia), Dominique Arrouays (ITPS/France), Svyatoslav Balyuk 

(Ukraine), Marta Camps Arbestain (ITPS/New Zealand), Laziza Gafurova (Uzbekistan), Hakki Emrah Erdogan 

(Turkey), Elena Havlicek (Switzerland), Maria Konyushkova (Russia), Ramazan Kuziev (Uzbekistan), Marc 

van Liedekerke (EC), Vitaliy Medvedev (Ukraine), Luca Montanarella (ITPS/EC), Panos Panagos (EC), Manuela 

Ravina da Silva (EC), Bülent Sönmez (Turkey). 


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in Europe and Eurasia


The majority of reports on the global state of soil degradation regard the European region as less disturbed 

compared with the situation in other regions. According to an ISRIC estimation (Oldeman, 1998), the 

average cumulative loss of productivity during the post-Second World War period due to human-induced soil 

degradation was estimated as 7.9 percent while in Africa it was 25 percent, and in Central America it was as high 

as 36.8 percent. However, the extent of soil degradation in Europe appears to be underestimated, because soil 

degradation on the territory of the European region has many facets, not all considered in previous estimates. 

The processes of human-induced soil degradation started in many parts of the region in ancient times, 

because many centres of agrarian civilization emerged in Europe and Eurasia several millennia ago: Greece, 

Anatolia and the Amu Darya delta are just the most remarkable examples. Since that time the pressure on the 

land has increased because of growing populations and the intensive migration of people due to a decline in 

natural resources and climatic fluctuations. The western part of the European region in comparison to other 

regions of the world has a history of over 200 years of industrialization which have placed additional pressures 

on the soil. 

Soil changes can occur naturally but are now under increasing threat from a wide range of anthropogenic 

pressures. Today these pressures represent the main reason for soil degradation in many parts of Europe. 

Soil resources are being over-exploited, degraded and irreversibly lost due to poor management practices, 

industrial activities and land-use changes. These issues in the region threaten soil’s key role as the basis for 

provision of food, feed, fibre and energy as well as for ecosystem services and mitigation of climate change. 

Knowledge on the state of soil resources in the region is good because of the generally high development 

of soil science and soil monitoring in the countries of the region. Nonetheless, an overview of the state of soil 

resources and of the development of land degradation for the whole region remains difficult because of the 

lack of harmonization of data, which were often obtained at different times using different methodologies 

(Jones and Montanarella, 2003; Morvan et al., 2008). 

In this chapter, we focus on anthropogenic degradation, e.g. alteration of soil properties induced by human 

activities that leads to declines in soil productivity and ecosystem services. The human activities in question 

include improper agricultural use, and soil disturbance and contamination due to urbanization, industrial and 

mining activities. 

11.2 | Stratification of the region 

The European region as considered in this report includes Europe sensu stricto plus Turkey and Eurasia. This 

larger definition extending beyond Europe proper entails consideration of a wider diversity of bioclimatic and 

soil resources and consequently of land use. The importance of agriculture varies among the countries of the 

region (Table 11.1). In terms of percentage area under agricultural use, the five leaders are Kazakhstan, Moldova, 

United Kingdom, Ukraine and Turkmenistan and the five least agricultural countries are Greenland, Norway, 

Finland, Sweden, and the Russian Federation. For the countries with the largest agricultural area, it should 

be noted that the figures require interpretation. For example, the high percentage of agricultural lands in 

Kazakhstan does not mean that the country has the highest pressure on natural ecosystems, because almost 

90 percent of its agricultural area is in fact occupied by rangelands. The countries with the least percentage 

of agricultural lands are the coldest countries of the region, where bioclimatic condition do not allow the 

extension of agricultural activities. 


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Area name 

Agricultural area of 

total land area, percent 

Area name 

Agricultural area of 

total land area, percent 

Kazakhstan 77.00 

Republic of Moldova 76.05 

United Kingdom 71.35 

Ukraine 71.32 

Turkmenistan 69.29 

Hungary 64.41 

Uzbekistan 63.04 

Denmark 62.82 

Ireland 62.79 

Romania 61.67 

Serbia 57.81 

Azerbaijan 57.54 

Spain 57.44 

Netherlands 57.12 

Armenia 56.21 

Kyrgyzstan 55.96 

Greece 55.38 

Czech Republic 55.09 

Serbia and Montenegro 54.81 

France 53.81 

Poland 53.17 

Turkey 52.27 

Luxembourg 49.95 

Italy 49.76 

Germany 48.59 

Bulgaria 47.95 

The former Yugoslav 

Republic of Macedonia 

46.78 

Belgium 45.57 

Belarus 44.20 

Lithuania 43.69 

Liechtenstein 42.44 

Slovakia 42.17 

Bosnia and 

Herzegovina 

42.07 

Andorra 41.81 

Albania 41.69 

Channel Islands 41.34 

Portugal 41.17 

Austria 39.59 

Georgia 38.34 

Montenegro 38.19 

Switzerland 38.11 

Tajikistan 33.47 

Malta 30.34 

Latvia 27.80 

Slovenia 24.62 

Croatia 21.98 

Estonia 20.45 

Iceland 18.43 

San Marino 16.67 

Cyprus 15.38 

Russian Federation 13.17 

Sweden 7.66 

Finland 7.47 

Norway 3.39 

Greenland 0.57 

Table 11.1 The percentage of agricultural land area of total land area in the countries of the European region Source: FAO, 2015. 


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The internal stratification within the region and within countries, including the extent of agricultural lands, 

depends mainly on bioclimatic conditions which determine agro-ecological zones. According to standard 

agro-ecological zoning (Fischer et al., 2002), the European region lies in the following agro-ecological zones: 

Arctic (only Greenland and Russia); Boreal continental (Eastern Russia); Boreal sub-continental (Russia, 

Scandinavian countries, Greenland); Boreal oceanic (Iceland and Greenland); Temperate continental (Russian 

and Kazakhstan); Temperate sub-continental (Eastern and Central Europe, Turkey, South Caucasian countries, 

Russia, Central Asian Countries); Temperate oceanic (Western, Central and Northern Europe); Sub-tropical 

with winter rainfall (Southern Europe, Turkey, South Caucasus); and Sub-tropical with summer rainfall (only 

small areas in Spain and southern France). 

The other approach for characterizing the internal stratification of the region is its division according to 

the major biomes. The digital map of terrestrial eco-regions presented below (Figure 11.1) delineates the major 

biomes found in the European region, based on the World Wildlife Fund’s eco-regions (The Nature Conservancy, 

2009). 

Produced by EU JRC 

Figure 11.1 Terrestrial eco-regions of the European region. Source: Olson et al., 2001. 

In brief, the relation of these major zones to agricultural development and soil degradation processes is the 

following: 

Polar and tundra, and taiga zone 

This zone represents a treeless polar ecosystem located in high latitudes in the European region in Russia 

and Scandinavia. The climate is characterized by long winters with months of total darkness and extremely 

frigid temperatures. Vegetation is mainly scattered, although sometimes it can be patchy, reflecting changes 

in soil and moisture gradients. Most precipitation falls in the form of snow during winter time. Soils tend to 

be acidic and saturated with water where not frozen. The region is sparsely populated, with agriculture in the 

tundra limited to reindeer grazing. Thus agricultural pressure on soils is not very strong; however huge areas 

are affected by mining and petroleum extraction. Some land degradation processes are triggered by humans 

indirectly. For example, in Siberia the processes of permafrost melting due to climatic change is resulting in the 

alteration of topography and causing severe damage to roads and buildings. Waterlogging is also a challenge 

in many areas. 


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Boreal forest/taiga 

These ecosystems cover extensive areas in northerly latitudes and with low annual temperatures. There are 

large expanses in central and eastern Russia, with medium precipitation of 40-100 cm yr -1 , partly in the form 

of snow. Predominant tree species are coniferous including Abies, Picea, Larix and Pinus as well as deciduous 

such as Betula spp. And Populus spp. The ground cover is mainly dominated by mosses and lichens. These 

biomes are known for slow regeneration of mature forests, due to the challenging climate and soil conditions. 

Forests are sensitive to acid rain and other forms of pollutant. Agriculture in the taiga is restricted to relatively 

small areas used for livestock and production of such crops as rye, flax, millet and vegetables. Some two 

decades ago, soil acidification induced by industrial contamination of the atmosphere, so called ‘acid rain’, 

was an important menace to soil quality in these areas. However, today the pressure of technogenic acid 

precipitation is considerably reduced (Jones et al., 2011). 

Broad-leaf and mixed forest zone 

This zone stretches across the European region from the British Isles to Western Siberia, and most of 

this territory is actually under cultivation. In the temperate climate, forests experience a wide variability in 

temperature and precipitation. Species such as oak (Quercus spp.), beech (Fagus spp.), birch (Betula spp.) 

and maple (Acer spp.) typify the composition of this biome. Structurally, these forests are characterized by 

four layers: a canopy composed of mature full-sized dominant species; a slightly lower layer of mature trees; a 

shrub layer; and an understory layer of grasses and other herbaceous plants. In contrast to tropical rain forests, 

most biodiversity is concentrated much closer to the forest floor. The zone has a favorable humid climate and 

soils with a relatively high natural productivity. Anthropogenic pressure is, however, strong due both to the 

intensive practice of agriculture and to high population density. The main threats to soils in this zone are water 

erosion favoured by intensive deforestation, and soil sealing and capping due to the high urbanization rate 

and dense infrastructure. In addition, the high degree of industrialization in this biome results in extensive 

contamination of the soils. 

Temperate coniferous forest 

This ecosystem is also known as ‘temperate evergreen forest’. It sustains the highest levels of biomass of 

any terrestrial ecosystem after tropical rainforest. The area has warm summers and cool winters, resulting 

in a high variation of the vegetation, e.g. needle leaf trees, broadleaf evergreen trees or a mix of both types. 

Temperate evergreen forests are common in the coastal areas of regions with mild winters and heavy rainfall, 

or inland in drier climates or hilly areas. Predominant tree species include pine, cedar, fir and redwood. This 

biome is mostly located in mountainous regions and the use of these areas is not very intensive. 

Temperate grassland zone 

This zone possesses the soils with the highest natural productivity such as Chernozems and Kastanozems. 

This high potential results in an intensive use of the land for agriculture which in places occupies up to 90-95 

percent of the total land area. The main threats to soils in this zone are water and wind erosion. These processes 

are the main reasons for the loss of organic carbon in soils; however, the loss of carbon by mineralization from 

arable lands is also a common process. Since the population density and the development of industry are high 

in this zone, soil sealing and capping and contamination are also threats. In places, especially in endorheic 

valleys, soil salinization may be observed. 


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Mediterranean zone 

This zone by definition is typical of the regions around the Mediterranean Sea. The ecosystems are 

characterized by hot and dry summers with cool and moist winters, with precipitation mostly during the 

winter months. Plant species are uniquely adapted to the stresses caused long, hot and dry summers. Most 

of the vegetation is adapted to fire and in fact depends on this disturbance for its sustainability. Some wildlife 

species undertake seasonal migration according to resource availability. The natural communities in this 

biome are highly sensitive to habitat fragmentation, grazing, and alteration of fire regimes (over-burning 

or fire suppression). Native species are also at risk from exotic species that easily establish and spread. To 

maintain the natural communities fire regimes need to be managed and exotic species controlled (WWF, 

2014). The ecosystems and soils are vulnerable due to the dryness of climate and to the abundance of shallow 

limestone soils. In many countries, transformation of pastures represents a serious problem. Erosion, organic 

carbon loss and decline in biodiversity are the main challenges for areas with Mediterranean climate. In places 

soil salinity may also limit the agricultural use of soils. 

Sub-arid and arid zone 

This zone includes deserts and semi- deserts, mostly located in Central Asia but also in some areas of 

Anatolia (Camci Çetin et al., 2007) and in the Iberian Peninsula. The vegetation in this zone is sparse, and all the 

ecosystems are subject to desertification. The main threats to soils in the zone are salinization, sodification, 

and wind erosion. Salinization is a natural process in many areas, also favored by initially saline parent material 

of marine origin. However, the most menacing process is irrigation-induced soil salinization, which leads to a 

drastic decline in soil fertility. 

11.3 | General threats to soils in the region 

Soil threats in the European region are complex, and although they are unevenly spread in the region, their 

dimension is continental and they are frequently inter-linked. If not managed, soil threats will lead to soil 

degradation, and the capacity of the soil to carry out its vital ecosystem functions will be lost. When many 

threats occur simultaneously, the combined effect tends to aggravate soil degradation (Jones et al., 2005). 

Climate change is likely to affect soil quality and cause land degradation through changes in soil moisture 

content (Calanca et al., 2006; Wong et al., 2011; García-Ruiz et al. 2011). For example, throughout central and 

northern Europe, evapotranspiration has increased by about 0.3 mm day -1 and this has the potential to deplete 

the normally adequate soil water store and limit plant growth. More frequent and severe droughts may cause 

plant cover to reduce leading to the onset of erosion and desertification (Jones et al., 2011). However, the 

precise impacts of climate change on soil degradation in Europe are still uncertain (Kovats et al., 2014). General 

threats to soils in the European region include the following. 

(1) Erosion by wind and water 

A recent report (Jones et al., 2011) estimated that in the 1990s 105 million ha, or 16 percent of Europe’s total 

land area (excluding Russia), were affected by water erosion, and that 42 million ha were affected by wind 

erosion. A new model of soil erosion by water constructed by the JRC has estimated the surface area affected 

in EU-27 at 1.3 million km², with almost 20 percent subject to soil loss in excess of 10 tonnes ha -1 yr -1 . In Russia 

the area affected by medium and strong water erosion is 51 million ha, 26 percent of the agricultural land 

area and about 3.5 percent of the total land area (Ministry of Natural Resources, 2006). In Ukraine the area 

affected by water and wind erosion is about one third of all agricultural land, or 14.4 million ha. In Moldova the 

area affected by water erosion is about 840 thousand ha, one third (33.6 percent) of the total area of arable 

lands in the republic (Leah, 2012). In Belorussia the area affected by water erosion is 467 thousand ha, and by 


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wind erosion 89 thousand ha; totally eroded lands cover about 10 percent of the territory of the country. In 

Turkey, where 80 percent of soils are located on slopes steeper than 15 percent, the area affected by moderate, 

severe and very severe erosion is 61.3 million ha, or 78.7 percent of the total area of the country; wind erosion is 

active on about 500 thousand ha (Senol and Bayramin, 2013). The area of eroded soils in South Caucasus varies 

between 35 and 43 percent of total agricultural lands, aggravated by the mountainous topography of the 

region. In Central Asia as a whole, the total area affected by water erosion is over 30 million ha, and by wind 

erosion about 67 million ha: in Uzbekistan up to 80 percent of agricultural land is affected by water erosion, 

and in Tajikistan the area of agricultural lands affected by water erosion is estimated by different sources 

at between 60 and 97 percent (CACILM, 2006). Long-term observations in Russia show that soil erosion on 

average decreases the yield of leguminous crops by 15 percent, of wheat by 32 percent, of potatoes by 45 

percent, and of perennial grasses by 25 percent (State Committee of Russian Federation on Land Resources 

and Land Planning, 1999). 

(2) Soil organic carbon change 

Organic matter is a key component of soil, controlling many vital functions (Jones et al., 2011). The loss of 

organic matter in soils is due both to erosion and to the increased rate of mineralization of organic carbon in 

arable soils. Methodologically, it is difficult to separate erosion and mineralization-driven loss of humus in 

soils. However, soils with negligible erosion loss commonly also lose organic carbon under cultivation. The 

rate of soil organic matter loss differs between mineral and peat soils. In the latter group, soil degradation 

after drainage may be fairly quick, producing an intensive flux of carbon dioxide to the atmosphere and in 

places leading to complete mineralization of the organic layer and exposure of infertile underlying sediments. 

This threat is discussed in detail in Section 11.4.3 below. 

(3) Soil contamination 

Due to more than 200years of industrialization, soil contamination is a widespread problem in Europe. 

The most frequent contaminants are heavy metals and mineral oil. The number of sites where potentially 

polluting activities have taken place now stands at approximately three million. The issue is discussed in detail 

below in Section 11.4.1. 

(4) Soil acidification 

Acidification involves the loss of base cations (e.g. calcium, magnesium, potassium, sodium) through 

leaching and their replacement by acidic compounds, mainly soluble aluminum and iron complexes. 

Acidification is always accompanied by a decrease in a soil’s capacity to neutralize acid, a process which is 

irreversible in nature except over very long periods. Regulatory controls initiated in recent decades to mitigate 

global warming have had a significant impact on the emissions of pollutants that cause acidification, mainly 

by decreasing SO 2 

emissions. By 2020, it is expected that the risk of ecosystem acidification will only be an 

issue in some hot spots, in particular in the border area between the Netherlands and Germany (EEA, 2010a). 

Recovery from acid deposition is characterised by decreased concentrations of sulphate, nitrate and aluminium 

in soils. An increase in pH and acid-neutralising capacity (ANC) coupled with higher concentrations of base 

cations would, in turn, improve the potential for biological recovery. However, given the delay in the response 

of soil to decreases in acid deposition, many decades are likely to be required for affected sites to recover fully. 

Additional information on trends in acidification is presented in the SOER 2010 Air Pollution Assessment (EEA, 

2010a). 


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(5) Salinization and sodification 

In Europe, salinization is generally the result of the accumulation of salts from irrigation water and fertilizers. 

High levels of salt eventually make soils unsuitable for plant growth. Improper irrigation and the use of highly 

mineralized irrigation water lead to rapid accumulation of soluble salts in soil profiles. This form of salinization 

affects approximately 3.8 million ha in Europe. Though a number of practices of saline soil reclamation exist, 

most of them are expensive and not very effective, and all of them are site-specific. This makes salinity control 

a challenging task. Geographically this threat is localized in the drier parts of the region, mostly in central Asia, 

in southern Russia, in Turkey, Azerbaijan, Greece, Hungary, and Spain. This threat is discussed in more detail 

in Section 11.4.4 below. 

(6) Waterlogging 

Waterlogging occurs in many soils affected either by a high groundwater table or by rainwater stagnation 

due to poor permeability. Most waterlogged soils have excessive moisture due to natural causes, but 

waterlogging may also be caused by improper irrigation practices or through disturbance of landscape 

hydrology by construction, mining and traffic activities. The Russian Federation is the country that has the 

most extensive area of waterlogged soils in the world. Excessively moist soils, both mineral and organic, 

occupy 360 million ha (21 percent of the total area of the country). Among the agricultural lands of Russia 

23.9 million ha or 10.1 percent are waterlogged. Unlike Russia, where the distribution of waterlogging is 

mostly due to natural reasons, in Central Asia excessive moisture is caused by irrigation. In Uzbekistan, the 

groundwater table is less than 2 meters below the surface in about one-third of irrigated lands. The area of 

waterlogged land varies from 40 percent in the Fergana Valley to 80 percent in downstream Amu Darya. In dry 

areas, waterlogging is commonly associated with salinization. 

(7) Nutrient imbalance 

The situation with the balance of nutrients in soils is much better in the European region than in most 

other parts of the world. However, there is considerable heterogeneity in the distribution of nutrients in soils. 

In Western Europe, the concentration of nutrients in soils is high due to application of high doses of fertilizer 

(Grizzetti, Bouraoui and Aloe, 2007). The absolute leader in the use of nitrogen fertilizers is the Netherlands, 

where in places the dose exceeds 170 kg of N per ha. Germany, France and the United Kingdom also apply 

nitrogen fertilizers intensively. The highest doses of phosphorus fertilizers – in doses higher than 21 kg P per ha 

– are used in some regions of Italy, Spain, France and Greece. These regions are running a risk of contaminating 

the ecosystem with excessive fertilizers (e.g. for France: Lemercier et al., 2008, Follain et al., 2009). In Eurasia, 

by contrast, the use of fertilizers is much lower, although Russia and Belorussia are major exporters of mineral 

fertilizers. The restricted use of fertilizers in these countries is in part due to the high natural productivity of 

their soils, but is also due in part to economic reasons. Underutilization of fertilizers in Central Asia is caused 

by the fact that small farmers cannot afford them or make money out of them. 

(8) Compaction 

Compaction can be induced by the use of heavy machinery in agriculture. Compaction reduces the capacity 

of soil to store and conduct water, makes it less permeable for plant roots and increases the risk of soil loss 

by water erosion. Estimates of areas of Europe at risk of soil compaction vary. Some authors estimate that 

36 percent of European subsoil has a high or very high susceptibility to compaction (Jones et al., 2011). Other 

sources report 32 percent of soils as being highly susceptible and 18 percent as being moderately affected 

(Jones et al., 2011). In Russia and Central Asia, soil compaction is also a challenge, especially where soils are 

naturally susceptible to compression, for example, in soils with vertic or natric properties, or in heavy textured 

soils. These kinds of soils are located mainly in the southern part of the Russian Federation, and in many places 

in Kazakhstan, Uzbekistan, Kyrgyzstan and Tajikistan. 


337 


(9) Sealing and capping 

Sealing occurs when agricultural or non-developed land is lost to urban sprawl, industrial development or 

transport infrastructure. It normally includes the removal of topsoil layers and leads to the loss of important soil 

functions, such as food production, water storage or temperature regulation. The population of the Europe is 

approximately 11 percent of the world population (740 million) and it has grown at a rate of 0.15 percent ayear 

during the last decade. The urbanization rate is high, particularly in the small and highly developed countries 

of Western Europe. It is the rapid and intensive expansion of urban and industrial development onto soils that 

makes soil sealing an important challenge for Europe. The issue is discussed in more detail in Section 11.4.2 

below. 

11.4. Major threats to soils in Europe and Eurasia 


Local contamination of soils is generally associated with intensive industrial activities, inadequate waste 

disposal, mining, military activities or accidents. Management of these contaminated sites is a tiered process 

starting with a preliminary survey (searching for sites that are likely to be contaminated), followed by site 

investigations where the actual extent of contamination and its environmental impacts are defined, and 

finally remedial and after-care measures. 

As discussed above (11.3.3), the number of sites where potentially polluting activities have taken place in 

Europe now stands at approximately three million. Due to improvements in data collection, the number of 

recorded sites is expected to grow as investigations continue. If current trends continue and no changes in 

legislation are made, the numbers reported above are expected to increase by 50 percent by 2025 (Jones et 

al., 2011; EEA, 2014). There is some evidence of progress in remediation of contaminated sites, although the 

rate is slow. In recent years, around 17 000 sites have already been treated while many industrial plants have 

attempted to change their production processes to generate less waste. In addition, most countries now have 

legislation to control industrial wastes and prevent accidents. In theory, this should limit the introduction of 

pollutants into the environment. However, recent events − such as the flooding of industrial sites in Germany 

during extreme weather events which led to the dispersal of organic pollutants, and the collapse of a dam at an 

aluminium plant in Hungary in October 2010 − show that soil contamination can still occur from potentially 

polluting sites. Trends in the deposition of heavy metals from industrial emissions are discussed in the SOER 

2010 Assessment on Air Pollution (EEA, 2010b). 

Diffuse soil contamination is one of the specific threats to soils in European region. Even though this form 

of pollution may be only barely apparent or may not even be directly apparent at all, it can cover very large 

areas. The contaminants include inorganic compounds such as metallic trace-elements and radionuclide, and 

organic compounds like natural and xenobiotic molecules. Radionuclides originate from nuclear accidents 

and nuclear tests, but there are also other sources such as fertilizer application (P). The most common 

xenobiotic chemicals include PAH, PCB and many pesticides still in use or inherited from the past (e.g. DDT 

and its metabolites). 

The economic feasibility of soil surveys is hindered by the diversity of contaminants – particularly of 

persistent organic pollutants (POPs), which are in constant evolution due to agrochemical developments – 

and by the transformation of organic compounds in soils by biological activity into diverse metabolites. The 

most widespread metallic trace elements in European soils comprise notably As, Cd, Cr, Cu, Hg, Ni, Pb, Zn, 

which are mobilized in one place but may then be transported to distant areas. 


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The real extent of diffuse soil contamination by metallic trace elements is not clearly known. Even though 

some EU member states and other countries have already implemented long-term soil surveys, they lack a 

harmonized soil monitoring system. 

The case study of Austria (Section 11.5.1) includes the description of: (i) local and diffuse pollution affecting 

Austrian soils; (ii) the effect of regulation on contamination trends; and (iii) the remediation activities being 

carried out. 


Soil sealing is especially intensive in Western Europe. On average, built-up and other manmade areas 

account for around 4 percent of the total area in EEA countries (data exclude Greece, Switzerland and the 

United Kingdom), but not all of this is actually sealed (EEA, 2009). EU member States with high sealing rates 

exceeding 5 percent of the national territory are Malta, the Netherlands, Belgium, Germany and Luxembourg 

(EC, 2011).The EEA has produced a high resolution soil sealing layer map for the whole of Europe for the year 

2006 based on the analysis of satellite images. Much more detail can be found in the SOER Assessments on 

the Urban Environment (EEA, 2010c) and Land Use (EEA, 2010d), as well as in EC (2011). 

Productive soil continues to be lost to urban sprawl and transport infrastructure. Between 1990 and 2000, 

the sealed area in the EU-15 increased by 6 percent and at least 275 ha of soil were lost per day in the EU, 

amounting to 1 000 km² per year. Between 2000 and 2006, the EU average loss increased by 3 percent, but by 

14 percent in Ireland and Cyprus, and by 15 percent in Spain (EC, 2011). A study by Huber et al. (2008) provides 

an interesting insight into the development of baselines and thresholds to monitor soil sealing. See also the 

SOER 2010 Assessment on Land Use (EEA, 2010b) for additional details on urbanisation. 

Soil sealing causes adverse effects on soil functions, or even their complete loss, and it prevents soil 

from fulfilling important ecological functions. Fluxes of gas, water and energy are reduced which affects, 

for example, soil biodiversity. The water retention capacity and groundwater recharge function of soil are 

reduced, resulting in several negative impacts such as a higher risk of floods. The reduction in the ability of soil 

to absorb rainfall leads to rapid flow of water from sealed surfaces to river channels, resulting in damaging 

flood peaks. Above-ground biodiversity is affected through fragmentation of habitats and the disruption of 

ecological corridors. These indirect impacts affect areas much larger than the sealed areas themselves. Builtup 

land is lost for other uses such as agriculture and forestry, as the soils which are sealed are often fertile and 

high value soils. Soil sealing appears to be almost irreversible and may result in an unnecessary loss of good 

quality soil. Soil sealing can lead to the contamination of soil and groundwater sources because of higher 

volumes of unfiltered runoff water from housing, roads and industrial sites. This is exacerbated during major 

flood events, as was demonstrated by the 2002 floods on the Elbe which deposited levels of dioxins, PCBs and 

mercury from industrial storage areas onto the soils of floodplains, well in excess of national health thresholds 

(Umlauf et al., 2005). 

In the Russian Federation the density of population is low and the area of settlements constitutes only 0.3 

percent of the national territory. However, the vicinities of megacities suffer intensive urban sprawl which 

takes land out of agriculture. 

11.4.3 | Soil organic matter decline 

Soil organic matter is essentially derived from residual plant and animal material, transformed (humified) 

by microbes and decomposed under the influence of temperature, moisture and ambient soil conditions. The 

stable fraction of organic substances in soil is known as complexed organic matter (previously referred as 

humus). Soil organic matter (SOM) plays a major role in maintaining soil functions because of its influence on 

soil structure and stability, water retention and soil biodiversity, and because it is a source of plant nutrients. 


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The primary constituent of SOM is soil organic carbon. Some 45 percent of soils in Europe have low or very 

low organic matter content (0–2 percent organic carbon). This is particularly evident in the soils of many 

southern European countries, but is also the case in parts of France, the United Kingdom, Germany, Norway 

and Belgium. A key driver is the conversion of woodland and grassland to arable crops. 

The soils of EU-27 Member States are estimated to store between 73 and 79 billion tonnes of carbon, which 

is equivalent to almost 50 times annual greenhouse gas emissions from the EU. However, intensive and 

continuous arable production may lead to a decline of soil organic matter. In 2009, European cropland emitted 

an average of 0.45 tonnes of CO 2 

per hectare, much of which resulted from land conversion (EEA, 2011). 

The conversion of peatlands and their use is particularly worrying. For instance, although only 8 percent 

of the farmland in Germany is on peatland, it is responsible for about 30 percent of the total greenhouse 

gas emissions of its whole farming sector (EC DG Environment and JRC, 2010). However, with appropriate 

management practices, soil organic matter can be maintained and even increased. Apart from peatlands, 

particular attention should be paid to the preservation of permanent pastures and the management of 

forests soils, as carbon age in the latter can be as high as 400-1 000 years (EC DG Environment and JRC, 

2010). Maintaining carbon stocks is essential for the fulfilment of the present and future emission reduction 

commitments of the EU. 

In Russia more than 56 million ha of mineral soils on agricultural land are characterized by the loss of organic 

matter (Shoba et al., 2010), in Ukraine, the area subject to loss of SOM is 18.4 million ha (Laktionova et al., 

2010), and in Moldova more than 1 million ha (Leah, 2012). Turkey was reported to be losing SOM from about 

70 percent of its agricultural soils (Senol and Bayramin, 2013), but the rate and extent of dehumification are 

unknown yet. In the South Caucasus republics, the loss of soil organic matter is not well documented, largely 

because in mountainous countries it is hard to separate out erosion and mineralization-driven loss of humus. 

A similar situation exists in the data reported for Central Asia: the cultivation of virgin lands in Kazakhstan 

resulted in the loss of approximately 570 million tonnes of carbon from soils, but a significant part has been 

transported by the wind. 

Several factors are responsible for a decline in SOM and many of them relate to human activity: conversion of 

grassland, forests and natural vegetation to arable land; deep ploughing of arable soils; drainage; fertiliser use; 

tillage of peat soils; crop rotations with reduced proportion of grasses; soil erosion; and wild fires (Kibblewhite 

et al., 2005). High soil temperatures and moist conditions accelerate soil respiration and thus increase CO 2 

emissions (Brito et al., 2005). Excess nitrogen in the soil from high fertiliser application rates and/or low plant 

uptake can cause an increase in mineralisation of organic carbon which, in turn, leads to an increased loss of 

carbon from soils. Maximum nitrogen values are reached in areas with high livestock populations, intensive 

fruit and vegetable cropping, or cereal production with imbalanced fertilisation practices. While in extreme 

situations, the surplus soil nitrogen can be as high as 300 kg N ha -1 (EC, 2002), estimates show that 15 percent 

of land in the EU-27 exhibits a surplus in excess of 40 kg N ha -1 . For reference, while rates vary from crop to crop, 

the IRENA Mineral Fertiliser Consumption Indicator (EEA, 2005a) estimates that average application rates of 

nitrogen fertiliser for EU-15 in 2000 ranged from 8 to 179 kg N ha -1 . 

There is growing realisation of the role of soil, in particular peat, as a store of carbon and recognition of soil’s 

role in managing terrestrial fluxes of atmospheric carbon dioxide (CO 2 

). Other than in tropical ecosystems, 

soil contains about twice as much organic carbon as above-ground vegetation. Soil organic carbon stocks 

in the EU-27 are estimated to be between 73 to 79 billion tonnes, of which about 50 percent is to be found 

in the peatlands and forest soils of Sweden, Finland and the United Kingdom (Schils et al., 2008). Peat soils 

contain the highest concentration of organic matter in all soils. Peatlands are currently under threat from 

unsustainable practices such as drainage, clearance for agriculture, fires, climate change and extraction. The 

current area of peatland in the EU is estimated at more than 318 000 km 2 , mainly in the northern latitudes. 


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While there is no harmonised exhaustive inventory of peat stocks in Europe, the CLIMSOIL report (Schils 

et al., 2008) estimated that more than 20 percent (65 000 km 2 ) of all peatlands have been drained for 

agriculture, 28 percent (almost 90 000 km 2 ) for forestry and 0.7 percent (2 273 km 2 ) for peat extraction. The 

degradation of organic soils is especially pronounced in Belorussia, where about 190 thousand ha of peat soils 

are strongly degraded: the peat layer was completely mineralized, and infertile sands were exposed on an area 

of 18.2 thousand ha. In the Russian Federation, the mineralization of peat has occurred on extensive drained 

areas with Histosols located mainly in the north of the European part of the country. The total area of drained 

peatland in Russia was estimated as 3.86 million ha (Inisheva, 2005), but the area of deeply degraded drained 

peat soils is unknown. Extensive areas of previously drained peatland have been abandoned as agriculture 

has shifted to more climatically favorable regions closer to markets. These dry peatlands are subject to fires in 

dry summer periods; in 2010 peat fires in Central European Russia caused an ecological catastrophe, driving 

millions of people from their homes and disrupting air transportation because of dense smoke over thousands 

of square kilometers. Another common practice in Russia is forest drainage, designed to improve forest 

productivity in waterlogged areas. Currently about 3 million ha of forested organic soils are drained in Russia, 

mainly in the European part. Though productivity has increased, the loss of organic carbon from soils is also 

evident. In Ireland, peatlands are widely used for energy production. Annual production of peat in the country 

peaked in 1995 at 8.0 million tonnes (Devlin and Talbot, 2014). 


While naturally saline soils exist in certain parts of Europe, the main concern is the increase in salt content in 

soils resulting from human interventions such as inappropriate irrigation practices, use of salt-rich irrigation 

water and/or poor drainage conditions. Locally, the use of salt for de-icing can also be a contributing factor. 

The primary method of controlling soil salinity is to use excess water to flush the salts from the soil. In most 

cases where salinization is a problem, this must inevitably be done with high quality irrigation water. 

Thresholds to define saline soils are highly specific and depend on the type of salt and land use practices 

(Huber et al., 2008). Excess levels of salts are believed to affect around 3.8 million ha in Europe (EEA, 1995). 

While naturally saline soils occur in Spain, Hungary, Greece and Bulgaria, artificially induced salinization is 

affecting significant parts of Sicily and the Ebro Valley in Spain and more locally in other parts of Italy, Hungary, 

Greece, Portugal, France and Slovakia (Figure 11.2). 

In the post-Soviet countries, saline soils cover the most extensive areas in Kazakhstan, Russia, Uzbekistan, 

Turkmenistan, Ukraine, and Azerbaijan (Figure 11.3). In other countries saline soils are present only locally, 

either on marine marsh deposits or in areas of salt mining. In Belorus, salt-affected soils occupy less than 1 

thousand hectares around the potassium fertilizer mines in Solikamsk region. The area occupied by saline soils 

in different countries is shown in Table 11.2. 


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Country 

Area of salt-affected soil 

(million ha) 

Data source 

Kazakhstan 111.5 Borovskii, 1982 

Russia 54.0 Shishov and Pankova, 2006 

Uzbekistan 20.8 Kuziev and Sektimenko, 2009 

Turkmenistan 14.1 Pankova, 1992 

Ukraine 4.0 Novikova, 2009 

Turkey 1.5 Senol and Bayramin, 2013 

Spain 0.63 Tóth et al., 2008 

Hungary 0.56 Tóth and Szendrei, 2006 

Azerbaijan 0.51 Ismayilov, 2013 

Table 11.2 The areas of saline soils in the countries with major extent of soil salinization in the European region 

Kazakhstan 

The area of saline soils in Kazakhstan, including Solonetz, alkaline soils, and complexes of Solonetz with 

other soils, is 111.55 million ha, or 41 percent of the national territory (Borovskii, 1982). Saline soils are present 

everywhere in the country except in mountainous areas. They are common in the steppe zone, where they 

cover about 30 percent of the area. In dry steppe, semi-desert and desert zones these soils occupy up to 50 

percent of the area. Salt-affected soils are represented mainly by Solonetz and alkaline soils. Solonchaks cover 

only 1–3 percent of the area of salt-affected soils in the steppe zone, and 7–13 percent of the area of salt-affected 

soils in the semi-desert and desert zones. 

Russian Federation 

The area of salt-affected soils in Russia, including Solonetz, alkaline soils and combinations of salt-affected 

soils with other soil groups, is 54 million ha or 3.3 percent of the total area of the country. The category of 

salt-affected soils includes all the soils with a salic horizon starting within the first 1 meter. Salt-affected soils 

in Russia are located in the European part of the country and in Eastern Siberia within the steppe, dry steppe 

and semi-desert zones. The major part of the area is occupied by Solonetz and alkaline soils. In Eastern Siberia 

and in the Far East, salt-affected soils are localized in closed inter-mountain basins with dry steppe landscapes 

and semiarid to arid climate. In Yakutia (Sakha), salt-affected soils form over permafrost in thermo-karst 

depressions (so-called alases) in taiga larch forest. Sporadically saline soils are present along the northern and 

eastern shores of the country, but their area is insignificant (Shishov and Pankova, 2006). 


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Prepared by I. Alyabina 

Figure 11.2 Soil salinization on the territory of the European region. Source: Afonin et al., 2008; Toth et al., 2008; GDRS, 1987. 

Uzbekistan 

The area of salt-affected soils in Uzbekistan is 20.8 million ha, or 46.5 percent of the country’s territory. 

Saline soils are everywhere in the country except in mountainous regions and well drained flood valleys. Saltaffected 

soils are represented by Solonchaks in closed depressions and the bottom of the largely dried-up Aral 

Sea, by Salic Calcisols in uplifted areas such as the Ustyurt plateau and the Kyzyl-Kum desert, by Takyric Salic 

Fluvisols in the alluvial and deltaic plains, and by extensive irrigated soils of various grades of human-induced 

salinization. 

Turkmenistan 

The area of salt-affected soils in Turkmenistan is 14.1 million ha or 28.7 percent of the total area of the 

country. These soils are distributed all over the country, but the major concentration is in the western part of 

the country close to the Caspian Sea. Most of these soils are not classified as Solonchaks and belong to the 

groups of Salic Calcisols and Takyric Salic Calcisols. 

Ukraine 

The area covered by salt-affected soils in Ukraine is about 4 million ha or 6.6 percent of the national territory 

(Novikova, 2009). Most of these lands are not used in agriculture; they are located largely in the southern and 

eastern parts of the country within the steppe and forested steppe zones. Most of these soils are classified as 

Solonetz and other soils with various grades of alkalinity. 


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Turkey 

The major areas with salt-affected soils in Turkey are: Konya-Eregli, the Aksaray and Malya plains of Central 

Anatolia, and the alluvial plains of lower Seyhan, Igdır, Menemen, Bafra, Söke, Acıpayam and Salihli. The 

distribution of the salt-affected arable lands is: 60 percent slightly saline, 19.6 percent saline, 0.4 percent 

alkali and 8 percent saline-alkali. Although sodium salts are the main components of the salt-affected soils, 

there are also magnesium soils in Denizli-Acıpayam, potassium-nitrate-alkali soils in Nigde- Bor, Kayseri, and 

gypsiferous soils in Central Anatolia. 

Spain 

In Spain 3 percent of the 3.5 million ha of irrigated land is severely affected, reducing markedly its agricultural 

potential, while another 15 percent is at serious risk. Soil salinization is a frequent problem in arid and semiarid 

regions like Southeast Spain (Hernández Bastida, Vela de Oro and Ortiz Silla, 2004). In these areas, demand for 

water for agriculture and increasing frequency of drought events have led farmers to irrigate with poor quality 

water. This has caused processes of soil degradation and salinization, limiting crop growth and impairing 

productive capacity (Pérez-Sirvent et al., 2003; Acosta et al., 2011). 

Hungary 

In Hungary Solonchak soils, which are by definition saline soils, occupy a total area of only 4.7 thousand 

ha. They are mainly located in low-lying areas, typically along the shorelines of saline/sodic lakes in the region 

between the Danube and Tisza Rivers, but they also occur in patches east of the Tisza River. These soils are not 

cropped, but sustain native halophyte vegetation which is grazed. Solonchak-Solonetz soils with a total area of 

65.9 thousand ha are also found largely between the Danube and Tisza Rivers, but above deeper groundwater 

levels of around one metre. These soils also sustain a native halophyte vegetation which is grazed. Mollic 

Gleyic Solonetz soils occupy a total area of 274.9 thousand ha and are characterized by large exchangeable 

sodium percent and a not very high salt content. Mollic Solonetz soils, with a total area 212.2 thousand ha, 

have only minor limitations for cultivation of crops and are typically under arable farming. 

Azerbaijan 

The area of salt-affected soils in Azerbaijan is estimated to be 510 thousand ha or 5.9 percent of the 

territory of the country. The saline soils are located mainly on the coastal plain of the Caspian Sea, in the Kura- 

Araksinskaya depression and in the Salyan, Mugansk, and Milsk plains. These soils are represented mainly by 

Solonchaks, Salic Gleysols and, in rare cases, by Salic Calcisols. 

While several studies show that salinization levels in soils in countries such as Spain, Greece and Hungary 

are increasing (De Paz et al., 2004), systematic data on trends across Europe are not available. 


11.5.1 | Case study: Austria 

Austria is a relatively small country that is land-locked in central Europe and shares borders with eight 

countries (Gentile et al., 2009). Austria’s location in the middle of Europe gives rise to specific environmental 

issues such as the pressures from intensive freight transit traffic (e.g. air emissions, habitat disruption) and 

the trans-boundary exchange of acidifying air pollutants and tropospheric ozone precursors (e.g. damage to 

forests and soil). In addition, only 37 percent of the national territory is suitable for permanent settlements. 


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This is due to the country’s geo-morphological conditions with more than 60 percent of the territory occupied 

by mountains. As a consequence, urban sprawl and land consumption occurs in restricted areas, with resulting 

high pressures on the environment. 

Overall, the Austrian soils are in a good condition, but their ecological functionality is at risk from diffuse 

and local accumulation of pollutants, intensive use of land, sealing and erosion. The more affected areas are 

located in the Alpine region, where forest soils are threatened by air deposition, and in urban areas, where 

sealing and contamination put urban soils under a growing pressure. Sealing is also present in rural areas due 

to the increasing urban sprawl and new road construction. In arable land, According environmental measures 

within Cross Compliance in arable land a treatment of the soil is prohibited, if the soil is water-satisfied. Soil 

erosion is higher on steeper slopes and land under permanent crops (vineyards, orchards), as well as on land 

cultivated with maize, sugar beet, potatoes and vegetables. 

The use of heavy machinery, especially in case of wet soil conditions, often results in the compaction of the 

topsoil, which can reach in some cases the subsoil layer. Soil compaction mainly occurs in areas with intensive 

agriculture and, locally, in other areas due to forest management activities. According environmental measures 

within Cross Compliance in arable land a treatment of the soil is prohibited, if the soil is water-satisfied. 

Floods happen occasionally in the floodplains of eastern Austria after extraordinary weather conditions 

(heavy rainfalls) whereas landslides occur quite often in the alpine regions with steep slopes. Adequate 

management measures for the protection of forests and afforestation, as well as technical engineering 

measures, are being implemented to prevent the consequences of such events and reduce the risks. 

Salinization is found in the areas around the lake Neusiedl, which is located in the north-eastern part, at 

the Hungarian border. Salinization is causing problems only in small areas with intensively managed and 

irrigated agricultural soils. Information on decline of organic matter is scarce. Some areas of arable land show 

low organic matter content. The evaluation of the Austrian environment programme for agriculture (ÖPUL) 

concerning organic matter contents in soil showed positive trends. 

Contamination from heavy metals is mainly due to long-range trans-boundary air pollution, especially in 

forest soils due to the high filter capacity of vegetation cover and the barrier effects of the Alps to air mass 

circulation. Contamination by heavy metals and persistent organic pollutants can also be found in restricted 

areas originating from different sources, e.g. local industrial sources, traffic especially near bigger population 

centres or agricultural sources. 

The restructuring of the industrial sector and in particular the decline of the heavy industry in the 1990s 

did not have observable effects on environmental pressures. Despite the decrease of the overall production, 

the contribution of this sector to the overall emissions is still considerable. Adverse effects of soil degradation 

are still to be expected, despite the continuing improvement in the implementation of regulations and the 

reduction of pollutant emissions, since many pollutants (e.g. heavy metals) are accumulating in the soil. The 

major indirect impacts are on biodiversity and the quality of groundwater resources. 


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The threats to soil in Austria 

Contamination 

Diffuse contamination 

Soil surveys targeted at the four most relevant heavy metals (mercury, lead, cadmium and copper) showed 

increased lead and cadmium concentrations in topsoil with respect to background values in the regions of the 

Northern and Southern Limestone Alps. This may be attributed both to local sources of pollution and to longrange 

trans-boundary air pollution. Lead enrichment is particularly high in grassland and forest soils – the 

latter due to the high filtering effect of the vegetation cover. The guidance values for lead established by the 

Austrian Standard (Önorm, 1975, 2004 1 ) were exceeded in more than 5 percent of monitored grassland sites, 

and in more than 3 percent of forest sites. Cadmium concentrations exceeded the guidance value in 5 percent 

of the monitoring sites in forests and in 6 percent of the sites in grassland areas. 

On the other hand, soil pollution with mercury and copper only occurs in restricted areas. In particular, 

copper contamination can be found mostly in the surroundings of industrial sites processing copper ore and 

in areas with intensive animal husbandry. The latter is due to the application of high amounts of pig manure 

with high copper content, of which the source is copper-enriched ready-made food. Other sources of copper 

inputs in soil are sewage sludge and compost as well as the application of pesticides containing copper. About 

2 percent of the forest and grassland monitored sites exceeded the guidance value for copper. 

Contamination from Persistent Organic Pollutants (POPs) was found in a limited number of sites, some 

of which required to be cleaned-up. POP pollution mainly occurs in urban areas and around industries. It can 

also derive from long-range trans-boundary air pollution. Emissions of POPs have been substantially reduced 

in the past years. This should have resulted in lower concentrations in the soil. However, a systematic survey 

targeted at organic pollutants in soil has not been carried out. For this and other reasons, such as the low 

mobility of these pollutants and the appearance on the market of new chemical products, the importance of 

POP pollution may be expected to increase in the future (Umweltbundesamt, 2004, 2007b). 

Contamination from local sources 

In Austria, soil contamination requiring clean up may be present at 2 500 sites. Potentially polluting 

activities are estimated to have occurred at 80 000 sites (including the 2 500 sites already mentioned) and 

investigation is needed to establish whether remediation is required. Approximately 70 heavily contaminated 

sites have been cleaned up in the past two decades. Industrial production and commercial services, municipal 

and industrial waste treatment, and oil storage are reported to be the most important sources of heavy 

contamination. National reports indicate that heavy metals, polycyclic aromatic hydrocarbons, cyanides 

and mineral oil are the most frequent soil contaminants at investigated sites. Nearly two-thirds of the 

remediation expenditure come from public budgets. Although considerable efforts have been made already, 

it will take decades to clean up the legacy of contamination. New contamination is not expected due to the 

implementation of prevention measures in place. 

Salinization 

In Austria, salinization is of little relevance as compared to other soil threats. According to an agricultural 

soil mapping survey carried out in the period 1958-1970, the areas where soil is affected by salts amounted 

to only 2 500 ha. Conditions for potential salinization do occur in small areas in Eastern Austria. These are 

areas with a negative water balance, salt-sensitive soils, a low groundwater table and salty groundwater. 

1 An ÖNORM standard is a national standard published by the Austrian Standards Institute. ÖNORMs are voluntary standards drawn up in committees of the 

Austrian Standards Institute 


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Future changes in climate and land management practices could lead to the salinization of soils also in these 

areas. In addition, soda-containing soils are estimated to cover only 2 000 ha. In these areas, strict rules for 

agricultural production apply. 

Erosion 

Soil erosion in Austria is mainly increased by unsustainable agricultural practices, construction of buildings 

and roads, and the use of leisure infrastructures. National estimates report that about 13 percent of the 

agricultural land or more than 5 percent of the total territory is potentially under a high risk for water erosion. 

The spatial distribution of potential erosion risk is very heterogeneous. The most affected areas include the 

productive areas of the southeast and northeast plains and hills, the Alpine foreland and the Carinthian basin 

(Strauss and Klaghofer, 2006). 

Except for the results of some scientific studies, Information on wind erosion is scarce. Loss of soil by wind 

has been observed in the lowlands of Eastern Austria. Areas at risk are sandy soils and, in the dry season, some 

areas covered with black soils (chernozems). In the past, some measures, such as reforestation of lowlands, 

were carried out to protect soil against wind. New windbreaks are planted annually, thus increasing the 

protected areas by several thousand hectares per year. The presence of wind erosion risk in sandy areas has 

been acknowledged since the 18th century. This early recognition of the problem and the measures adopted 

have resulted in the stabilisation of erosion in these areas (Strauss and Klaghofer, 2006). 

Soil erosion is not expected to increase in the future due to the implementation of prevention and reduction 

measures such as the measures included in the national Agri-Environmental Programme. In 2008 these 

measures reduced the soil erosion rate by 3 to 18% depending on the region, in average by 10 percent (AGES, 

2011). However, major pressures may come from future climate and land use changes (conversion of grassland 

into arable land) or significant changes in crop rotation, although these are not very likely to occur. 

Decline in soil organic matter 

According to the results of a recent monitoring programme that measured the content of organic carbon 

in topsoil, more than half of all grassland and forest sites in Austria have a content of organic matter in topsoil 

of over 8 percent, which is comparatively high by global standards. On the other hand, in arable land, most of 

the monitored sites show an organic matter content ranging from 2 percent to 4 percent (medium by global 

standards), while a low or very low content (< 2 percent) is found in a quarter of the sites. In the areas with low 

organic matter content, the natural soil functions will be at risk in the long run (Umweltbundesamt, 2004). 

Measurements of organic matter content in topsoil also provide an indication of the content in soil organic 

carbon. 

Overall, the organic carbon stock in Austrian soils is not expected to decline in the future, due to the 

implementation of soil organic matter preservation measures in agriculture. Such measures of the national 

Agri-Environmental Programme have increased the humus content of arable soils from 1991-1995 until 2006- 

2009 by 0.1 up to 0.4 percent depending on region (AGES, 2011). 

Sealing 

Following the general trend in Europe, the sealing of the soil due to the increase of built-up areas and 

transport infrastructure has shown a growing trend in Austria in the past decades, although population has 

increased only slowly. 


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In total, about 7 percent (or around 5 500 km²) of the country area is occupied by buildings or transport 

infrastructures, and about 56 percent of this area is sealed. About half of the new residential buildings in 2001 

were single family dwellings or semi-detached houses which, in comparison to multi-family residences or 

other high-density structures, occupy a considerably larger surface area (Umweltbundesamt, 2006). In the 

period 2012-2014, the average increase in built-up areas was about 19 ha day -1 . This resulted in a daily increment 

of soil actually sealed of 10 ha in 2012 and 2014. This figure still exceeds by a factor eight the relevant policy 

target (Umweltbundesamt, 2007a). These high rates may lead to the saturation of the available space in some 

regions. In Vorarlberg, for example, 29 percent of the permanent settlement area is already built-up. 

These increases are due to changes in the standard of living and in lifestyles and to the development of 

associated settlement and transport activities, rather than to population growth. This is particularly evident 

in rural regions where the built-up area continues to grow despite a net decrease in population. 

Hydro-geological risks 

Erosion and erosion control have been a major issue for a long time in Austria, due to the country’s specific 

geo-morphological configuration. More than 60 percent of the territory is occupied by mountains. The focus 

of past and current activities is on the control of torrents and avalanches, as these are major threats to human 

life in alpine environments (Strauss and Klaghofer, 2006). 

According to BMLFUW (Austrian Federal Ministry of Agriculture, Forestry, Environment and Water 

Managemen 2 ), about 67 percent of the territory may be classified as either part of a torrent watershed, 

avalanche watershed or a general risk area. The regional coverage ranges from 16 percent in Burgenland to 

91 percent in Tyrol. The amount of budget available for measures against these risks increased from 70 million 

EUR in 2001 to 148 million EUR in 2009. 

In case of extraordinary weather conditions (heavy rainfalls), floods happen occasionally in the floodplains 

of eastern Austria. The flood events in August 2002 affected large parts of the national territory. Particularly 

Upper Austria and Lower Austria suffered heavy damage, as floods reached areas that were previously 

considered as safe. More details on this event can be found in the special chapter on floods of the 7th national 

State of Environment report (Umweltbundesamt, 2004). 

Cross-cutting issues 

Brownfields 

In general, the remediation of contaminated sites based on fit-for-use remediation goals should be seen not 

only as bringing an improvement to the status of the environment through the restoration of soil functions but 

also as bringing benefits to economy and society. In Austria, there is a potential for brownfield redevelopment. 

The average consumption of green-field areas for housing and traffic was 7 ha day -1 in 2014. On the other hand, 

only 37 percent of the national territory is suitable for permanent settlements. 

According to an unofficial definition, brownfields are sites of formerly industrial or commercial land, now 

derelict or underused, or sites that have been affected by former uses of the site or surrounding land. The 

latter may require intervention before they can be returned to beneficial use, particularly where there are 

contamination problems. The number of brownfield sites in Austria is in the range of 3 000-6 000, covering 

an area between 8 000 and 13 000 ha. According to estimates based on their previous use, about 85 percent 

of the industrial brownfield sites may present no or little contamination problem and could be revitalised and 

reused without public funding for remediation. 

2 http://www.bmlfuw.gv.at/en.html 


348 


Considering an increase of industrial brownfield sites of about 3 ha per day, about a quarter of the annual 

land requirement for housing and economic activities could be saved by reconverting brownfields to a 

productive use. Some measures have been proposed to this end. These include policy measures, sustainable 

and innovative land management, and mechanisms for the involvement of stakeholder groups. However, 

redevelopment activities are yet to be started. For more details see (Umweltbundesamt, 2007b). 

Soil services 

The main soil services in Austria include: 

• protection of groundwater and spring water in mountain areas, resulting in the availability of water 

in sufficient quality and quantities (about 99 percent of drinking water supplies originate from 

groundwater and spring water) 

• high diversity and mosaic distribution of geology and soils, which enables a high level of diversity of 

landscapes and biodiversity 

• availability of highly productive soils for agricultural and forestry production 

• availability of soils for building purposes, although this function is limited due to the relatively small 

permanent settlement area available (37 percent of the total area of the country) 

• The main impacts of soil degradation in Austria are: 

• biodiversity decline from soil sealing and contamination 

• impairment of groundwater quality from diffuse pollution and local contamination - there may be 

2 500 sites in Austria needing remediation, of which less 3 percent have been cleaned up since 1989 

• destruction of natural landscapes from soil sealing and unsustainable agricultural practices 

Hot spots 

The Alpine region is an environmentally sensitive area with a high level of diversity of landscapes, soils, 

flora and fauna. This region is under threat of acidification and contamination from deposition of air-borne 

pollutants on the one hand, and from erosion and landslides because of its steep slopes on the other hand. 

Some industrial areas are seriously contaminated from diffuse sources. These include, in particular, the city 

of Linz, the Inn valley in the Tyrol, and Arnoldstein in Carinthia. In addition, a high concentration of sites which 

are potentially contaminated can be found in the most urbanised and industrialised regions, in particular the 

cities of Vienna, Linz and Salzburg, the Inn valley in the Tyrol and the Mur and Mürz valley in Styria. 

Outlook 

Soil resources in Austria are on average in a good condition; however soil functions are still being threatened 

by the deposition of airborne pollutants, by a legacy of contamination in industrial and urban areas, and by the 

continuing increase in the built-up area. 

There are some uncertainties on future trends of soil contamination due to the lack of data on the presence 

of organic pollutants in soil and the appearance on the market of new chemical products whose effects on 

the environment are not fully understood (Umweltbundesamt, 2004 and 2007b). However, the inputs of 

pollutants (in particular lead, cadmium and POPs) in the soil are expected to decrease, since emissions and 

thus depositions are diminishing due to the implementation of regulations and preventive measures in place. 

On the other hand, acidifying substances, in particular NOx from traffic sources, are expected to increase. 

Moreover, soil contamination and its adverse effects are still to be expected in the long run since many 

pollutants (e.g. heavy metals and POPs) have low mobility and high persistence and accumulate in the soil. In 

addition, the increase of the emissions of acidifying substances may result in an increase of the pressures on 

forests and forest soils. The major indirect impacts will be in terms of the loss of biodiversity and the quality 

of groundwater resources. 


349 


In urban and industrial areas, no new contamination is expected due to prevention measures in place. 

Nevertheless, the clean-up of historical contamination will continue to pose a challenge. In fact, although 

considerable efforts have been made already, in particular in the investigation and monitoring of contamination, 

a slow progress is made in the implementation of remedial measures. According to the national vision for 

contaminated sites the identification of all historically contaminated sites and the remediation measures at 

heavily contaminated sites shall be completed until 2050 (BMLFUW, 2008). 

The amount of soil actually sealed through the construction of buildings and infrastructures is currently 

increasing at a rate of 10 ha day -1 . This figure exceeds ten times the national 2010 sustainability target. As in 

Austria only 37 percent of the territory is suitable for permanent settlements, high increases of built-up areas 

may also lead to the saturation of the available space. This is more likely to occur in some regions currently 

registering the highest rates, especially in rural areas. On the other hand, the redevelopment of brownfields 

and the clean-up of historical contamination are expected to provide opportunities for the reduction of the 

consumption of green-land, as well as opportunities for economic and technological development, and job 

creation. Brownfields in Austria could cover about one quarter of the current needs for land. 

Soil degradation in agricultural areas, especially erosion, compaction and decline in organic matter content 

is not expected to increase in the future due to the implementation of prevention and reduction measures such 

as notably those measures included in the national Agri-Environmental Programme. For the same reasons, 

the organic carbon stock in Austrian soils is expected to remain stable on average. 

Climate change and the development of the tourist sector may result in increased hydro-geological risks. 

Investigations and monitoring of historical contamination are quite advanced. More is known on 

contaminated sites but remediation activities are progressing slowly. If current trends are maintained, it will 

take decades to clean-up a legacy of contamination. In Austria, remediation is aimed at removing the source 

of contamination and restoring the soil functions to a certain extent. The objective is to make the soil again 

fit for specific uses, in particular for the protection of groundwater resources (ultimately, the source for about 

50 percent of drinking water supplies). Further progress with the clean-up of historical contamination and 

brownfield redevelopment will also provide opportunities for the reduction of land consumption, economic 

and technological development, and job creation. 

11.5.2 | Case study: Ukraine 

About 60 percent of the soils of Ukraine are Chernozems – soils known for their unique structure, chemical 

properties and inherent fertility. These soils are distinguished by a very deep (more than 1 m) humus-enriched 

layer, perfectly expressed granular structure, almost optimal bulk density, and a good and satisfactory stock 

of nutrients. However, these favourable soil properties are only present in soils under virgin ecosystems. 

Chernozems are the best soils in the world (the ‘tsar of soil’) according to Vasiliy Dokuchaev, the founder of 

modern soil science. However, are very sensitive to anthropogenic intervention. In particular, when intensively 

tilled, they rapidly degrade. As a result, in recentyears these soils have been characterized in Ukraine by a 

significant decline of their potential productivity, equal to 80-90 million tonnes of grain annually, or up to 2 

tonnes per head of the population. It has proved almost impossible to maintain good ecological conditions for 

these soils. 

The fragility of the soils of Ukraine is well known and has been the subject of many studies. Nevertheless, 

this has not deterred intensive development which has led to their severe degradation. About a third of arable 

land is eroded, 30 percent of organic matter has been lost, approximately 40 percent of soils have a compacted 

layer, and stocks of nutrients have noticeably decreased. Where soils have been improved, numerous problems 

have also been observed (Balyuk and Medvedev, 2012). 


350 


Comparison of cultivated soils with virgin analogues shows that for the last 40-50 years the most typical 

processes were (Bulygin, Breus and Seminozenko, 1998; Grekov et al., 2011; Medvedev, 2012): 

• Loss of humus in arable soils: at the end of the 1980s, 0.5-1.5 tonnes per ha were being lost annually. 

Between 2005 and 2009 the rate of organic matter loss was 0.42-0.51 tonnes per ha annually. 

• Increasing deficiency of labile nutrients, especially nitrogen (declining from -41.4 kg per ha in 2001 to – 

56.4 kg per ha in 2009) and potassium (declining from -32.9 to – 64.2 kg per ha between 2001 and 2009). 

• Acidification of Chernozems, especially in the Cherkassy and Sumy regions, where the drop in pH was 

0.3-0.5 units. 

• Compaction, particularly in the western forested steppe but widespread on 40 percent of arable land 

nationwide. Compaction is characterized by a destruction of structure, lumpiness and crusting. 

• Reduction of the depth of upper layer of the soil due to erosion, reaching several centimeters in 

Chernozems according to modelled data, and also in the drained soils of Polissiya. 

• Secondary alkalinization and salinization of irrigated soils, accompanied by a reduction of peat depth. 

Among other negative processes of local importance are: contamination with radionuclides and heavy 

metals; waterlogging; flooding; iron, calcium carbonate and aluminum accumulation; desertification; and 

alkalinization and soda formation. 

The types and extent of degradation of arable soils in Ukraine are shown in Table 11.3 and Figure 11.3. The 

estimation of soil degradation has been carried out using the technique proposed by van Lynden (1997). The 

sources of data have been: (i) the results of the agrichemical certification of fields conducted every five years 

since 1965; and (ii) the database of the National Scientific Center (the ‘O.N. Sokolovsky Institute for Soil Science 

and Agrochemistry Research’). The database has provided information on the morphological, physical, 

physicochemical and chemical properties of more than 2 500 soil profiles, and also information from long 

field experience on tillage and application of fertilizers (Laktionova et al., 2010; Grekov et al., 2011). 

Soil degradation in Ukraine is mainly the result of the use of inappropriate farming technology. Chernozems 

are vulnerable to mechanical deformation due to their low bulk density before tillage in the spring, and also 

to the influence of moisture owing to the low stability of the swelling smectite minerals which predominate 

in their mineralogical composition. 

The problem has been aggravated because state and regional programmes of soil protection slowed down 

significantly after 1991. These programmes had obtained important results in protecting and restoring soils up 

to the end of the 1980s but during the last two decades measures aimed at improving soil fertility have been 

significantly reduced. 

Soil degradation is a major problem in Ukraine. There is little realization of the threat it represents for the 

present and especially future generations. Issues include the absence of effective mechanisms to enforce laws 

on soil protection, and unbalanced and insecure land tenure. Combating soil degradation requires raising 

awareness at all levels of society, wide educational activity, active dissemination of knowledge, and gradual 

formation of a new attitude to soil resources. 


351 


Type of degradation 

Extent, 

percent of 

arable land 

Fertility decline and reduced humus content 43.2 

Compaction 38.2 

Sealing and crusting 37.5 

Water erosion, surface wash 16.8 

Soil acidification 14.1 

Waterlogging 12.9 

Soil pollution by radionuclides 10.9 

Wind erosion: loss of topsoil 10.5 

Soil contamination with pesticides and other organic contaminants 9.2 

Soil contamination with heavy metals 8.0 

Salinization / alkalinization 4.3 

Water erosion: terrain deformation by gullying 2.6 

Off-site effects of water erosion 2.5 

Lowering of the soil surface 0.4 

Wind erosion: terrain deformation 0.4 

Desertification 0.2 

Table 11.3 Types and extent of soil degradation in Ukraine 


352 


Figure 11.3 Some types and extent of soil degradation in Ukraine. Source: Medvedev, 2012. 

11.5.3 | Case study: Uzbekistan 

Uzbekistan is one of the flattest countries of Central Asia. About 80 percent of the national territory 

consists of plains, with mountains located in the extreme east of the republic. The climate of Uzbekistan is 

‘continental dry’, with significant daily and seasonal fluctuations in air temperature. The summer is long and 

hot, the autumn is relatively wet, and the weather in winter is variable. The arid climate favors desertification 

and maintains relict accumulations of soluble salts in soils and sediments. 

The entire area of the country is 44.9 million ha, and agricultural lands constitute 46.1 percent of the national 

territory. The distribution of soils in Uzbekistan reflects a complex system of pedo-geographical regularities 

(Figure 11.4). The westernmost part of the republic is occupied by a desert zone that can be subdivided into 

a sub-boreal desert (Central Kazakhstan) and a subtropical desert (Turan) desert. The boundary between 

the two types of desert corresponds to the northern limit of possible cotton cultivation. In the lower belt of 

the piedmont subtropical semi-desert there occur mainly Calcisols which are replaced as elevation increases 

by mountainous Cambisols under steppe and forest vegetation. Anthropogenic factors strongly modify the 

morphology and pedogenesis of the soils. Irrigated soils are mainly transformed into Anthrosols - some of 

the soils in the Amu Darya delta have been cultivated and irrigated for more than three thousandyears. The 

country is vulnerable to negative environmental impacts due to its natural climatic conditions. Irrigated 

agriculture is localized in the plain and piedmont parts of the republic and is characterized by varying levels of 

technology and intensity and by the varying quality of the land. 


353 


The natural drivers of land degradation and desertification are the following: 

• Climatic characteristics, such as aridity, continentality, wind action etc., which cause such phenomena 

as drought, hot winds, deflation, and atmospheric transportation of sand, salts and dust; 

• Topography, with slopes favouring the development of water erosion and landslides, and flat areas 

with depressions creating conditions for waterlogging and salt accumulation. Topography also favours 

the formation of specific intensive winds which play an important role in wind erosion; 

• Parent material, whose peculiarities are reflected in the soil profile (texture, gypsum and salts content) 

and which also determine the susceptibility of soils to wind erosion, karst phenomena, and the soil 

buffering capacity; and 

• Extreme natural phenomena, such as forest and grassland fires, floods etc. which affect the 

development of soils. 

Anthropogenic factors affect land resources and trigger degradation processes in the following ways: 

• Irrigation without a proper drainage system and inappropriate regulation of the collector-drainage 

water lead to salinization and waterlogging; 

• Inappropriate use of pastures leads to overgrazing, formation of exposed soil surface and destruction 

of the soil structure and, as a result, to the development of deflation under the effect of wind and high 

temperature; and 

• Forestry strategy allows excessive logging which causes soil erosion on the slopes in mountainous 

areas, soil deflation and the expansion of sands on the fertile lands on the plains (Arabov, 2014). 

Other economic activities such as industry, municipal and domestic wastes, transport emissions, and 

unreclaimed mining spoil also contribute to land degradation. The natural processes of land degradation 

and desertification are slow; their effect becomes evident only after decades or centuries. However, human 

activities accelerate these natural processes, and the results of anthropogenic degradation processes appear 

in a short period of time. 

Box 1 | The catastrophe of the Aral Sea 

An illustrative example of the menacing scale of ecological and socioeconomic disasters caused by 

inappropriate use of natural resources is the catastrophe of the Aral Sea. Its volume has been reduced more 

than 13 times, and its area by more than seven times. The primary cause being the diversion of inflowing 

rivers for irrigation projects. The shoreline has moved hundreds of kilometers. Salt concentrations have 

reached 120 g per liter in the western part and 280 g per liter in the eastern part of the sea (Arabov, 2014). 

The sea has split and is now on the verge of extinction: only two small components separated by a dam are 

left, the deeper western part, and the ‘Small Aral’ in Kazakhstan, Most of the shore is surrounded by a rind 

of salt cover over marshy clays and sands. 

The processes of environmental change in the region have combined with global climatic changes and 

resulted in the intensification of seasonal droughts. The Aral catastrophe has aggravated the continentality 

of climate, increasing dryness and temperature in summer and prolonging cold and severe winters. In the 

Aral region, the number of days with the temperature over 40°C doubled, while in the rest of Uzbekistan 

it has gone up by about 1.5 times. According to expert evaluation, by 2035-2050 the air temperature in 

the region may increase by a further 1.5-3.0°C. On the dried surface of the sea bottom, there are extensive 

white salt crusts, covered in places by wind-blown sand. This territory forms a new desert called ‘Aral Kum’ 

that covers 5 million ha. This Aral Kum desert, which is still growing, has already absorbed 2 million ha 

of arable lands and led to degradation of pastures, riparian forest and other vegetation. Satellite images 

illustrate the penetration of plumes of salts and dust from the Aral Kum for 8 001 000 km, deep into 

densely populated zones. 


354 


The plains in the basins of Amu Darya and Syr Darya are lowlands with no natural drainage. Due to the dry 

climate, the low precipitation and the strong evaporation, these plains act as accumulators of easily soluble 

salts in the topsoil. For this reason, the development of irrigated agriculture, starting from the piedmont 

areas, requires careful attention to current or historic salt accumulation in the sediments. Farmers also have 

to be aware of the danger of secondary salinization. Salt-affected soils currently constitute more than 46 

percent of the total irrigated area in Uzbekistan, including moderately saline soils (25 percent of the total 

area), and strongly saline soils (over 6 percent). The worst affected areas are the regions of Karakalpakstan, 

Bukhara, Khorezm, Dzhizak, Syrdarya, Andijan, Kashkadarya, Navoi, Samarkand, Surkhandarya. Some 

districts in Tashkent and Fergana regions are also affected. In the Samarkand region, the prevailing type of 

salinity is magnesium-carbonate salts accumulation. Salinization of some newly irrigated lands is followed 

by the formation of gypsum-enriched soils that are difficult to reclaim. Gypsum layers and horizons impede 

water infiltration and decrease the efficiency of leaching doses designed to wash salts from the soil profile. The 

total area of gypsum-containing soils is about 350 000 ha. 

Figure 11.4 Soil map and soil degradation extent in Uzbekistan. Source: Arabov, 2010. 

Across Uzbekistan, all types of erosion can be found: surface runoff and irrigation erosion, destructive 

mudflows, wind erosion, and direct negative effects of wind on plants. Wind erosion and negative wind effect 

on plants affect 21.4 million ha or 80 percent of all agricultural lands (Figure 11.5). Of the 3.7 million ha of irrigated 

lands, three quarters – 2.8 million ha – are eroded to various extents. Agricultural lands also suffer from water 

and irrigation erosion. Moderately and strongly eroded soils constitute 12 percent of the agricultural land pool 

and about 5 percent of the irrigated land pool (Kurbanov, 2001). 

The major part of the country is occupied by pastures that cover an area of 20.6 million ha and serve as 

the main source of fodder for livestock (Arabov, 2014). Rational use and protection of the soils of pastures 

are issues in natural resources conservation and use. The state of pastures is currently endangered. During 

the last 70-80years, the soils of pastures suffered a drastic decrease in humus and nutrient content and they 

have been affected by salinization and by water and wind erosion. Other negative processes include soil 

compaction, alkalinization, and decline in biological activity and resulting loss of soil 


355 


Thus, the key challenges in soil degradation in Uzbekistan are: (i) secondary salinization of irrigated lands; (ii) 

waterlogging of irrigated agricultural lands; (iii) depletion of soils, including the loss of humus and nutrients; 

(iv) compaction; (v) surface runoff and irrigation erosion in mountainous and piedmont areas; (vi) deflation 

and pasture degradation in desert regions where transhumance is practiced; (vii) deforestation and loss of 

biological diversity; (viii) soil contamination with agrochemicals and industrial waste; (ix) desertification in 

the regions bordering the bottom of the dried-up Aral Sea; (x) inappropriate methods of land management; (xi) 

poor crop management (lacking or wrong crop rotation, in places insufficient or excessive use of fertilizers); 

(xii) insufficient irrigation; and (xiii) breakdown of the rules governing sustainable management of pastures 

(Gafurova et al., 2012). 

Clearly the protection of land and soils and their sustainable economic use are huge challenges for 

Uzbekistan. Article 55 of the Constitution reads: “The earth, minerals, water, flora and fauna, and other 

natural resources are national wealth, requiring their rational use and protection of the State”. In order to 

provide comprehensive rehabilitation, conservation, protection and improvement of soils and their fertility, 

and to improve overall environmental conditions, the following steps are required: 

• develop agricultural techniques aimed at restoring and enhancing soil fertility, including approaches 

to land reclamation and the promotion of farming practices which improve the physical, chemical and 

ecological status of the soil 

• develop rapid methods of large scale soil mapping and automate the process of compiling maps and 

soil assessments, using remote sensing and GIS technologies 

• develop techniques for preventing the processes of surface, gully and wind erosion, for predicting soil 

erodibility, and for recovering and improving the fertility of eroded soils 

• develop approaches to stopping processes of soil salinization and introduce more effective and 

innovative ways of desalinization and reclamation of saline soils 

• conduct targeted research to establish the levels and pathways of soil contamination by various 

toxicants, including fluorine, heavy metals, pesticides and others, and develop measures to prevent 

soil pollution by these substances 

• develop integrated science-based recommendations for the assessment of soil fertility of both arable 

lands and pastures to promote sustainable economic land use in the republic 

11.6 | Conclusion 

The inherent complexity and spatial variability of soil makes the evaluation of the impact of any change 

difficult. Transformations of features such as texture and mineralogical composition will only occur over 

geological time spans while properties such as pH, organic matter content or microbial activity will show 

a more rapid reaction. In addition, the response of a particular soil type may be both positive and negative 

depending on the function in question. For example, rising temperatures and precipitation may support 

increased agricultural productivity on soils previously deemed marginal, but such a transformation can lead to 

a deterioration of soil biological diversity and an increased risk of erosion. Quantitative assessments of future 

trends in soil characteristics and properties are limited. As a consequence, this chapter provides an outlook 

only for a selected number of issues. Considerably more effort is required to model changes in the state of soil 

conditions in relation to drivers such as changes in land use and climate. 

Based on the above finding, an assessment is made of the status and trend of the ten soil threats in order 

of importance for the region. At the same time an indication is given of the reliability of these estimates 

(Table 11.4). 


356 


Threat to 

soil function 


and land take 

Salinization 

and sodification 

Contamination 

Organic 

carbon change 

Nutrient 

imbalance 


Loss of soil 

biodiversity 

Summary 

In densely populated 

Western Europe soil 

sealing is one of the most 

threatening phenomena. 

Salinization is a 

widespread threat in 

Central Asia, and it is 

challenging in some areas 

in Spain, Hungary, Turkey, 

and Russia. 

Soil contamination is a 

widespread problem in 

Europe. The most frequent 

contaminants are heavy 

metals and mineral oil. 

The situation is improving 

in most regions. 

The loss of organic 

carbon is evident in most 

agricultural soils. Peatland 

drainage in northern 

countries also leads to 

rapid organic carbon loss. 

In Russia, extensive areas 

of agricultural lands were 

abandoned that resulted 

in quick organic mater 

accumulation; however, 

some of these areas 

are now again used for 

agriculture. 

In the western part of the 

region the loss of nutrients 

is compensated by 

application of high doses 

of fertilizers. In the eastern 

part the use of fertilizers 

is insufficient, and in most 

soils nutrient mining 

results in intensive mineral 

weathering. 

Water erosion is active 

in all the cultivated 

mountainous and rolling 

areas; the worst situation 

is observed in Turkey, 

Tajikistan and Kyrgyzstan. 

Due to the attention 

paid to this threat it is 

controlled in most areas, 

especially in the EU. 

Loss of biodiversity 

is expected in the 

most urbanized and 

contaminated areas of the 

region. However, there 

are almost no qualitative 

estimations of the 

biodiversity loss in soils. 


Confidence 



357 



acidification 

Waterlogging 

Compaction 

Acidification due to acid 

rain was a challenge in 

Northern and Western 

Europe. The situation is 

now improving, though 

several decades will be 

needed for complete soil 

recovery. 

Waterlogging is mostly 

associated with 

irrigation in Central Asian 

countries. Most cultivated 

irrigated soils there 

are waterlogged. This 

phenomena in Central Asia 

is commonly associated 

with salinization. 

The use of heavy 

machinery and 

overgrazing are 

threatening in almost all 

the agricultural areas. 

Table 11.4 Summary of soil threats status, trends and uncertainties in Europe and Eurasia 


358 


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12 | Regional assessment 

of soil changes in Latin America 

and the Caribbean 

Regional Coordinator: Maria de Lourdes Mendonça-Santos (ITPS/Brazil) 

Regional Lead Author: Juan Comerma (Venezuela) 

Contributing Authors: Julio Alegre (ITPS/Peru), Ildefonso Pla Sentis (Spain), Carlos Cruz Gaistardo (Mexico), 

Rodrigo Vargas (Mexico), Diego Tassinari (Brazil), Moacir de Souza Dias Junior (Brazil), Sebastián Santayana 

Vela (Peru), Maria Laura Corso (Argentina), Vanina Pietragalla (Argentina), María Nery Urquiza Rodríguez 

(Cuba), Candelario Alemán García (Cuba), Segundo Sacramento Urquiaga Caballero (Peru/Brazil), Maria de 

Lourdes Mendonça-Santos (ITPS/Brazil), Miguel Taboada (ITPS/Argentina), Olegario Muniz (Cuba), Carlos 

Henriquez (ITPS/Costa Rica) and David Espinosa (ITPS/Mexico). 


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in Latin America and the Caribbean


This chapter discusses the status of soil resources in Latin America and the Caribbean (LAC). The LAC 

region, as defined by FAO, includes the following countries: Antigua and Barbuda, Argentina, Bahamas, Barbados, 

Belize, Bolivia, Brazil, Chile, Colombia, Costa Rica, Cuba, Dominica, Dominican Republic, Ecuador, El Salvador, Grenada, 

Guatemala, Guyana, Haiti, Honduras, Jamaica, Mexico, Nicaragua, Panama, Paraguay, Peru, Saint Lucia, Saint Kitts and 

Nevis, Saint Vincent and the Grenadines, Suriname, Trinidad and Tobago, Uruguay and Venezuela. The emphasis will 

be on human-induced ‘anthropogenic’ changes, and not on natural causes, although it is not always easy to 

separate them. 

The LAC region extends from Mexico at latitude 32 degrees north to Tierra del Fuego at 55 degrees south. 

It also includes the Caribbean islands. The presence of 12 out of 14 terrestrial biomes (Olson et al., 2001), the 

rugged relief in Central America and along the western side of South America, and the large lowlands in central 

South America that also include interior wetlands, all combine to make LAC the most bio-diverse region in the 

world. In fact, eight of the 17 mega-diverse countries of the world are located in Latin America: Bolivia, Brazil, 

Colombia, Costa Rica, Ecuador, Mexico, Peru, and Venezuela. 

In terms of natural resources, Latin America is one of the richest regions of the world. With only 8 percent of 

the population, it has 23 percent of the world’s potential cropland, 12 percent of the actually cultivated land, 46 

percent of the globe’s tropical forest and 31 percent of the planet’s fresh water (Garret, 1997). The region could 

provide a further 800 million ha of land for agriculture (Laegreid, Bockman and Kaarstad, 1999). However, 

most of these potential areas are under tropical rainforest and clearing would cause severe environmental 

changes with dramatic effects on many ecosystem functions. Agricultural conversion of natural ecosystems 

(grass-shrub-savannas and forest) as a percentage in LAC is of the order of 30 percent, representing slightly 

over 600 million ha of agro ecosystems, a figure similar to Africa but smaller than Europe and much smaller 

than Asia. Compared to other regions of the world, these converted agro-ecosystems had a medium to low 

intensity use of fertilizers and irrigation for much of the 20th century (UNDP, 2000). However, use of fertilizer 

and irrigation increased dramatically in recent decades as agriculture expanded onto the temperate and 

subtropical plains (Grau and Aide, 2008; Viglizzo and Jobbagy, 2010). 

Another important characteristic of this region is that agriculture started in the mountains, mainly because 

of the presence of serious diseases in the lowlands, notably malaria. Consequently, the sloping lands of the 

Andes and the Central America Mountains have been cultivated the longest, starting with the Incas and Mayas. 

Mountainous areas in LAC still have high populations that practice both intensive agriculture like horticulture 

and more extensive land uses like pasture and coffee growing. Exploitation of the flatter lowlands such as the 

Pampas and Cerrados, with tropical, subtropical and temperate climates, started only more recently. This is 

where most of the present intensive farming takes place, including the use of fertilizers and machinery in the 

production of cereals, legumes and other crops (Viglizzo and Jobbagy, 2010). 

The main soil threats are related to natural features of physiography and to the type of vegetation cover. 

Anthropic and cultural features also play an important role, especially inappropriate agricultural practices 

which are a consequence of inequitable and insecure land tenure, insufficient research and lack of extension 

services. 

Water erosion and landslides are prominent threats in the sloping lands of the mountains, especially when 

the slopes have been burned and overgrazed. Loss of soil carbon mostly occurs after deforestation and change 

of land use to permanent grassland. In semiarid and arid areas where irrigation is applied, salinity and sodicity 

are important threats. In areas with more intense land use and the employment of heavy farm machinery, 

compaction occurs. Also in these areas, the use of amendments, fertilizers and other agrochemicals bring the 

threats of nutrient imbalance, acidification, contamination and loss of biodiversity. In addition to induced 


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flooding on rice fields, waterlogging and flooding are on the rise in the region as a consequence of the 

combination of higher rainfall promoted by the phenomenon of ENSO (the El Niño Southern Oscillation), and 

by land use changes which decrease ground cover by perennial vegetation. Acidification also occurs, mostly in 

deltas where the drainage of marine clays produces acid sulphate soils. 

Overall, the most important ecosystems services affected in LAC are: (1) climate regulation, through the 

carbon and nitrogen cycles, especially due to the immense deforestation rate up to 2004 of the humid tropical 

forests, mostly in the Amazon Basin; (2) water regulation, through changes in quantity and quality of water 

production in the mountains, which is also due to deforestation of sloping lands accompanied by strong water 

erosion and landslides; and (3) loss of biodiversity, another ecosystem service threatened by deforestation and 

change in land use/land cover (Viglizzo and Frank, 2006; Gardi et al., 2014). 

12.2 | Biomes, ecoregions and general soil threats in the region. 

In order to give information on soil, land use and ecosystem services affected by soil threats, this chapter 

uses the 'Biomes’ (environmental zones with similar climate, fauna and flora) outlined in the Soil Atlas of 

Latin America and the Caribbean (Gardi et al., 2014, Figure 12.1). The boundaries of the Biomes correlate with 

major soil-landscapes. We will describe and discuss the Biomes in sequence starting from the most extensive. 

However, this sequence does not necessarily follow the significance and gravity of soil threats. 


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Figure 12.1 Biomes in Latin America and the Caribbean. Source: Olson et al., 2001. 

Tropical and Subtropical Moist Broadleaf Forest is by far the most extensive Biome in LAC. Essentially it 

is a permanent very humid, high temperature forest. The landscape goes from flat plains to rolling slopes. The 

soils are dominated by medium to highly weathered Acrisols, Ferralsols and Plinthosols, which are in general 

acid and unfertile. This Biome has the greatest biodiversity of plant and animals and also shows little resilience 

to human intervention. It spreads from the eastern coasts of Mexico and Central America, through part of the 

Caribbean islands, the Pacific Coast of Colombia and the Atlantic coast of Guyana and Venezuela, down to 

most of the Amazon basin in Ecuador, Peru and Brazil, and finally to the western and southern coast of Brazil. 

Because this Biome has had the largest deforestation rate in LAC (FAO, 2005, 2010) and because this 

practice removes a large and continuous addition of organic matter to the soil, the loss of land cover and 

soil organic carbon is clearly the most important threat to ecosystem function. Another threat in deforested 

areas is water erosion, which occurs when heavy rains fall on bare soil, resulting in major erosion problems 

on the widespread gentle and moderate slopes. As most of the nutrients in these soils are contained in the 


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organic matter of the topsoil, its removal and rapid decomposition affects soil biodiversity, and in the long run 

nutrient imbalances may appear. Other ecosystem services affected include: the carbon and nitrogen cycles 

and their contribution to climate regulation; water quality and quantity; and landscape stability. 

Tropical and Subtropical Grasslands, Savannahs and Shrub lands are the second largest Biome in LAC. 

The climate is characterized by alternating wet and dry periods and by high temperatures. The vegetation 

consists predominantly of grasses with different densities of trees. In general, the topography is flat to 

gently undulating. For the most part, and especially in Brazil (Cerrados), eastern Venezuela and Colombia, 

the dominant soils are acid lowfertility Acrisols, Ferralsols and Arenosols, while in the younger surfaces, the 

dominant soils are more fertile Luvisols, Phaeozems and Vertisols. 

In these areas, the principal crops are annuals such as corn, soybean, sorghum, beans, cassava and cotton, 

grasses like Brachiarias, and legumes like Stylosanthes. In recentyears, sugarcane and other biofuel crops have 

also been planted (Miyake et al., 2012). Brazil, with the largest area of savannahs in LAC (around 200 million 

ha), has 55 million ha of introduced pastures and 22 million ha under annual crops (Sousa, 2011). In general, the 

soils in this Biome are low in organic matter, very infertile, and acid throughout. Counteracting the acidity and 

infertility of these soil conditions requires the use of amendments including gypsum and limestone, fertilizers 

such as rock phosphate, and inoculants. Erosion has been the main threat, but following major research and 

extension programmes, conservation tillage is increasing and is having an impact on the problem. Another 

threat has been compaction, not only by farm machinery but also by overgrazing. In grasslands where minimal 

fertilizer is used, a problem of nutrient imbalance has been reported (Guimarães, 2013). In agricultural land 

too, there is a growing imbalance of nutrients, especially in recentyears with the intensification of agriculture, 

where fertilizer use has fallen well short of the demand of the crops (Urquiaga et al., 2014). This is a threat 

throughout LAC. However, inputs of fertilizer have increased in recent years. 

Ecosystem services are affected both positively and negatively in this Biome. The rise in food production is 

the main positive effect. However, there are significant negative effects too: the carbon and nitrogen cycles 

are affected by the higher rate of organic matter decomposition; there is loss of water quality due to erosion 

and sediment movement; and there are biodiversity losses (Viglizzo and Frank, 2006; Viglizzo and Jobbagy, 

2010). 

Deserts and Xeric Shrublands occupy the third place amongst LAC Biomes in terms of area. They are 

characterized by low precipitation, high evaporation and quite windy conditions. The temperatures are 

variable: in the tropics, for example in northern Brazil, the coastal area of Peru and the Caribbean coast of 

Venezuela and Colombia, temperatures are quite warm; in northern and central Mexico and Chile, by contrast, 

a cold period occurs. In all cases, vegetation is dominated by cactuses and thorny shrubs that are very sparse 

and resistant to drought. The most typical soils are calcareous or gypsiferous, shallow, saline or sodic and 

reflect the very limited amount of leaching. Calcisols, Gypsisols, Arenosols, Regosols and Solonchaks are 

common soil groups. 

Because of its aridity, this Biome requires irrigation for most forms of agricultural development. Irrigation 

requires not only a source of water, but also water of good quality and a good drainage system to leach the 

soluble salts common in this Biome. Preventing salinization is difficult but restoring salinized soils is even 

harder. The use of waste water from cities may bring additional problems, particularly contamination with 

heavy metals. Due to the sparse vegetation coverage and the presence of strong winds, it is common for wind 

erosion to increase after human intervention. 

The principal ecosystem service that may be affected is the productivity of the soils due to salinization. 

Water quality may worsen for the same reason. 


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Temperate Grasslands, Savannahs and Shrublands have been widely studied and documented by various 

authors (Paruelo, Guerschman and Verón, 2005; Satorre, 2005; Viglizzo and Frank, 2006; Álvarez et al., 2009; 

Lavado and Taboada, 2009; Viglizzo and Jobbágy, 2010). This Biome is predominantly located in the Argentinian 

Pampas. Its central plains are dominated by grasses on flat to gently sloping lands, with a temperate climate, 

and rains ranging from 1 500 mm in the northeast to 400 mm in the southwest. These areas have some of the 

most fertile soils of the world, the Phaeozems, although more than 13 million ha with natural saline-sodic soils 

(e.g. Solonetzs) also appear in this Biome. 

Despite the wide adoption of no-tillage, intensive annual cultivation (largely of soybean) and the lack 

of rotation with other crops or pastures have resulted in soil degradation by wind and water erosion, 

waterlogging, compaction, sealing/capping, and soil fertility depletion (Satorre, 2005; Lavado and Taboada, 

2009; Alvarez et al., 2009; Sainz Rozas et al., 2011; Viglizzo and Jobbagy, 2010). Clearing of forested lowlands 

to produce annual crops (soybean, cotton etc.) has also led to salinization or sodification in areas where the 

groundwater table has risen (Paruelo, Guerschman and Verón, 2005; Viglizzo and Jobbagy, 2010). 

Dry Tropical and Subtropical Broadleaf Forest occurs in many different zones in LAC. These zones share 

common characteristics: they are all situated below 1 000 m elevation, all experience high temperatures, and 

all have at least one dry season when the trees lose their leaves. These forests are more common in hilly and 

mountainous landscapes with soils of medium fertility such as Luvisols, and Cambisols. Large populations 

live in these areas and deforestation and annual burning are common. These forests occur in western Mexico, 

Costa Rica, Cuba, the north of Venezuela and Colombia, the coasts of Ecuador and Peru, and central and the 

north east of Brazil. 

This Biome is very attractive for agricultural development. The soils are fertile and the climatic conditions 

are favourable for human habitation and for the growth of many crops, including corn, beans, potatoes, 

sugarcane, fruits and coffee. Large areas of grassland are used for extensive pasturing. Deforestation is the 

major threat, and is in practice irreversible as the dry period makes it difficult for natural regrowth to occur. 

As the accumulation of organic matter in these soils is medium or low, deforestation also brings the threat 

of organic carbon loss, and the increase of soil temperature in bare soils accelerates decomposition. Steeper 

slopes are hard to cultivate and here the most common land use after deforestation is extensive pasture. After 

a few years, soil compaction develops, particularly along the small terraces where the animals pass. This 

compaction also reduces rainfall infiltration and consequently increases water erosion. 

Ecosystem services affected are principally the carbon and nitrogen cycles and climate regulation due to the 

decrease in organic matter. Water production and quality may also decrease due to compaction by overgrazing 

and farm machinery which increases erosion. As erosion reaches high levels in many slopes, landscape stability 

can also be affected. 

Montane Grasslands and Shrublands are present at high altitudes throughout the Andes, but mostly in 

Peru and Bolivia. Locally they are called Punas or Paramos. They occur above 3 000 m or even higher, and are 

dominated by grasses and small shrubs. Temperatures during the day may be quite high but frosts can occur at 

night. In general these area are quite dry. The topography may be mountainous or large mesas or high plains. 

Soils are generally rich in organic matter, but rather shallow. Predominant groups are Regosols, Leptosols and, 

in the drier zones, Solonchaks. 

Most of the land remains bare or is used for extensive pastures. Most of the pastures are natural, but some 

have been introduced and have adapted to the conditions. In some parts of LAC, especially in northern South 

America, some areas are also used for intensive horticultural crops, including potatoes, carrots and quinoa. 

In these cases, conservation practices, irrigation and high inputs of organic and inorganic fertilizers are used 

throughout the year (Comerma, Larralde and Soriano, 1971). Although there are no data on impacts, it is 


369 


suspected that contamination of soils and water from excessive use of fertilizers and other agrochemicals may 

be occurring. This can also affect soil biodiversity, which may be very important given the unique vegetation 

and fauna of this Biome. 

This Biome is of high importance for certain ecosystem services, notably water production. It is located at 

the top of many watersheds and is a continuous source of pristine water, in many cases related to the process 

of thawing. Many soils in the Biome are rich in organic matter in the topsoil. Careful attention has to be paid 

to C and N cycles because of the services provided. The soil gene population, due to its unique nature, should 

also be studied and protected. 

Tropical and Subtropical Coniferous Forests occur at high and medium altitudes, mostly in Mexico and 

south to Nicaragua. Small patches also exist in the Dominican Republic, southern Brazil and along the Chile- 

Argentine border. These forests occur mostly in mountainous landscapes and on many different geologic 

materials, including volcanic ashes. Most of the soils are Umbrisols, Luvisols, Leptosols and Andosols. The 

vegetation is dominated by many types of conifer with a diverse understory. 

Deforestation for the establishment of pastures and wood harvesting are major land uses. In many cases 

these land uses are accompanied by burning. Deforestation and burning create the threats of reduced organic 

carbon and of water erosion. The main ecosystem services affected are water production/ regulation, C and N 

cycling, and landscape stability. 

Temperate Broadleaf and Mixed Forest is confined to a temperate permanent forest in southern Chile. The 

forest occurs mainly in valleys and on the slopes of high mountains. The climate is very humid, cold and with 

few variations during the year. Soils are shallow and mostly Alisols, Andosols, Cambisols, and Histosols. 

Because of the low temperatures and the rough relief, the land is largely a protected area, and only a few 

valleys are used for pasture. As human intervention is very low, threats and impacts for ecosystem services are 

very limited. 

Flooded Grasslands and Savannahs in LAC occur in tropical alluvial plains, mostly in Brazil, Bolivia and 

Paraguay. The Pantanal, which stretches across all three countries, is the world’s largest tropical wetland, and 

is a highly productive environment. The area is recognized by UNESCO as a World Natural Heritage Site and 

Biosphere Reserve. Other important areas occur in Venezuela and Colombia on flat alluvial plains. They all 

have in common the predominance of native pastures adapted to flooding, few trees at higher elevations, and 

a seasonal period of flooding alternating with a dry season. The predominant soils are Gleysols, Stagnosols, 

Vertisols, Plinthosols and Histosols. The most widespread use of this biome in LAC is as pasture for bovine 

and, more recently, bubaline cattle. Its strategic importance in the production system is in the supply of green 

pastures for the dry season. 

Economic development in the Pantanal region, especially on the plateau of the Rio Taquari Basin, 

has intensified the input of sediments to the Pantanal lowlands, causing serious social, economic and 

environmental impacts on the region (Galdino, Vieira and Pellegrin, 2006). 

The main ecosystem service is the provision of food and fiber, interacting with the service of water 

regulation. The ecosystem is unique and rich in flora and fauna. Soil biodiversity is of prime importance and 

should be investigated and protected. 

Mediterranean Forest and Shrubs occupy a small strip of Chile near the coast. The Biome has a warm 

temperate climate, dry during the warm period and rainy in the winter time. The vegetation, relief and soils 

are very heterogeneous, as they represent a transition between tropics and temperate, and between dry and 


370 


humid. The most common soil groups are Regosols, Leptosols and Andisols. The degree of human intervention 

is so complete that there are few remnants of the original vegetation. 

Because of its Mediterranean climate, the natural productivity of the area is very high. This has been utilized 

by Chileans to create very important agricultural development areas, especially in the valleys. Irrigation, 

tillage and large quantities of fertilizers and other agrochemicals are used to obtain high yields. Deforestation 

has resulted in significant erosion threats; high levels of input use have led to contamination; and in the drier 

zones salinization is a threat due to the quality of irrigation water. 

The ecosystem service positively affected is the provision of food. Negative effects are on water quality and 

possibly on human health, and loss of biodiversity related to contamination. 

Mangroves are the smallest Biome in LAC in terms of area. They occur along the coastlines of all LAC 

countries and are particularly associated with the deltas of important rivers like the Amazon and the Orinoco. 

They are located where there is the mixture of fresh and saline water. Mangroves provide many ecological 

services and are considered very fragile as disturbances can produce irreversible consequences. Soils are 

mostly Histosols, Gleysols and Acid Sulphate soils. 

Mangrove areas are mostly protected because they are hard to drain and the soils are poor, with 

predominance of reduced or organic soils. In cases where development projects have been implemented, the 

main land uses are pasture, rice, or plantation crops such as oil palm or bananas. Especially when mangroves 

are underlain by marine sediments, it is very common that after drainage, oxidation of Iron Sulphide (pyrite) 

will occur, producing extreme acidification and the formation of acid sulphate soils. This is an extreme case of 

the threat of acidification and reclamation is extremely difficult. After drainage, organic carbon is lost and the 

land surface subsides. 

The ecosystem service principally affected is the reduction of productivity, especially if acid sulphate soils 

are formed. The Carbon and Nitrogen cycles and water quality will also be affected. Cultural heritage will also 

be affected, as local tribes – for example, the Waraos in the Orinoco Delta – have been living on the natural 

products extracted from this ecosystem for more than 4000 years. 

12.3. General soil threats in the region 

In recent years, the onset of climate changes, notably more intense and concentrated rainfall events and 

higher evaporation, has begun to bring change in the pace and intensity of threats such as erosion, flooding 

and desertification (IPCC, 2014). At the same time, other threats to soils and ecosystem services have also 

increased, including threats to organic carbon, biodiversity, crop production, and water quantity and quality. 


This threat is considered one of the most important in the region, because it has an impact on very large 

populations, particularly those concentrated in the mountainous regions of the Andes, Central America, 

Mexico and the Caribbean. Water erosion and landslides occur mainly on steep slopes that have been deforested 

or in dry mountain areas which are used as pastures and which have been overgrazed. Water erosion also 

occurs on the gentler slopes of the cerrados as well as in parts of the pampas subject to intensive cultivation. 

Erosion and landslides remove fertile topsoil, affecting crop productivity, making tillage more difficult, and 

producing sediments that affect fields and infrastructure downstream and cause flooding in flat areas. The 

threat is considered in more detail in Section 12.4.1. 


371 



Changes in organic carbon occur mostly if carbon supplied by vegetation decreases, as happens with 

deforestation, or if mineralization is increased as happens with ploughing (Sánchez, 1981). LAC contains about 

half of the world’s tropical forest, and until recently, it had the highest rate of deforestation, which drastically 

reduced organic inputs to soils. The region also has some of the best soils in the world, for example in the 

Pampas. These soils are rich in organic matter content and very fertile, but continuous cropping has increased 

mineralization and reduced organic carbon. In both cases, soils are becoming less productive, limiting their 

ecological services (Lavado and Taboada, 2009; Viglizzo and Jobbagy, 2010; Gardi et al., 2014). A recent study 

(D’Accunto, Semmartin and Ghersa, 2014) found that uncropped agricultural borders are highly effective in the 

mitigation of soil organic carbon losses. The details of changes in SOC are outlined in Section 12.4.2. 


Natural or primary salinization and sodification are quite common in the arid and semiarid regions of 

LAC, including in Mexico, Cuba, northern South America, Peru, Northeast Brazil and southern Argentine. 

Human-induced threats are also present in these regions because irrigation is common and the quality of 

the water used and the lack of drainage induce salinization. Even though it may not occupy large areas, salt 

accumulation is an important threat because it severely reduces crop productivity. It is very difficult to prevent 

and even more difficult to reclaim soils once salinized. It has been estimated (AQUASTAT, 1997) that 18.4 million 

ha in LAC are affected by salinization caused by irrigation. The problem also appears in humid climates, where 

topsoil in large plains with high and saline groundwater and sodic soils (e.g. Solonetzes) may be salinized by 

upward soluble salt rises promoted by land clearing and overgrazing (Taboada, Rubio and Chaneton, 2011, 

Bandera, 2013, Di Bella et al., 2015). A country-by-country analysis is provided in Section 12.4.3. 


The largest proportion of soils in LAC are acid and have naturally low fertility. As a result, amendments and 

fertilizers are required for sustainable production. These inputs also serve to boost productivity where areas for 

agriculture and animal production are already limited and expansion of production requires intensification. In 

recentyears, contrasting cases of nutrient imbalance have emerged as a result of the high levels of N-fertilizers 

(ammonia) applied in high-input production systems (Espinosa and Molina, 1999). These N-fertilizers increase 

the acid ion sources in the soils, even in very fertile soils. Nutrient impoverishment has been documented in 

coffee and sugar cane plantations on Andisols, reducing yields (Bertsch et al., 2002). Nutrient imbalance could 

also appear in the highly fertile Pampas region, where farmers have historically applied low amount of fertilizers 

(Lavado and Taboada, 2009). The imbalance could be greater were it not for the significant contribution of 

biological nitrogen fixation (BNF) in agriculture and livestock production of the region (Urquiaga et al., 2012; 

Franzluebbers, Sawchika and Taboada, 2014). The contribution of BNF is particularly important for soybean 

in the Pampas and for legume pastures in the Southern Cone (Alves, Boddey and Urquiaga, 2003; Campillo et 

al., 2003, 2005). 


In LAC, conservation areas can provide a bank of original soil biodiversity (Alegre, Pashansi and Lavelle, 1996, 

White et al., 2005; Urquiaga et al., 2014). One study (Ferraro and Ghersa, 2007) found that the community 

of microarthopods was highly sensitive to crop management and resulting soil conditions. Microbiological 

indices and enzymatic activity are useful new indicators and have been used in different studies in the region 

(Balotta et al., 2004; Sicardi, Garcìa-Préchac and Frioni, 2004; Nogueira et al., 2006; Green et al., 2007; 

Franchini et al., 2007; de Moraes Sa and Lal, 2009; Romaniuk et al., 2012; Henríquez et al., 2014). 


372 



In LAC there are two main causes of compaction: livestock production and machinery transit. Overgrazing 

by cattle and sheep causes degradation of pastures and increased erosion as found in mountainous regions, 

the Cerrado, coffee plantations and the arid Patagonia (Bertiller, Ares and Bisigato, 2002; Henríquez et al., 

2011; Taboada, Rubio and Chaneton, 2011; Pais et al., 2013). The widespread use of farm machinery, especially 

in the Cerrados and the Pampas has caused shallow soil compaction and poor structural conditions in topsoil, 

especially associated with soybean mono-cropping and long winter fallow periods (Taboada et al., 1998; Botta, 

Tolon-Becerra and Melcon, 2009; Botta et al., 2010; Alvarez et al., 2014; Franzluebbers et al., 2014). 


Many flat areas are affected by episodic human-induced waterlogging and ponding. This can result from 

poor topsoil structural and drainage conditions which limit infiltration rates. It may also be associated 

with extreme rainfall events linked to the periodic climate phenomena of ‘El Niño’. Waterlogging has been 

documented in both agricultural and pasture soils throughout the region. There are increasing problems of 

catastrophic flooding, landslides and sedimentation (Pla, 2003, 1996a, 2011; Restrepo et al., 2006). 


Natural acidification is a common process in the soils of LAC and is very intense in tropical areas of the 

region, because of the high rainfall. Acidic soil parent materials are also widespread (Kämpf, Curi and Marques, 

2009; Fassbender and Bornemisza, 1987). As industrialization is not intensive and widespread in the region, 

soil pollution from industrial sources is not common. Anthropogenic acidification in soil could also appear 

where there is excessive N-fertilization on crops like banana, vegetables and oil palm and under intensive 

coffee systems (Sánchez, 1981; Espinosa and Molina, 1999). 


Different human activities may result in the pollution of soils and adjoining water bodies caused by fertilizers 

and agrochemicals used in high-input agriculture, and from mining and oil spills (Nriaugu, 1994; Malm, 1998; 

Mol et al., 2001). Residues of herbicides such as glyphosate have been observed in soils and groundwater in 

fields devoted to no-till farming (Ometo et al., 2000; Christoffoleti et al., 2008; Cerdeira et al., 2011; Aparicio 

et al., 2013). 

The use of mercury compounds in mining activities and the use of huge amounts of water for shale oil 

exploitation are causing downstream pollution in soils and waters (Nriaugu, 1994; Malm, 1998; Mol et al., 

2001). The increasing use of agricultural by-products and sludges increase N concentrations in groundwater 

and cause eutrophication of lagoons (Torri and Lavado, 2008). 

12.3.10 | Sealing 

According to the United Nations 1 , the population of Latin America is currently 630 million, 8.6 percent 

of the world’s population. This figure is expected to increase by 25 percent by 2050. Most Latin American 

countries are experiencing relatively low death rates and declining birth rates, resulting in slower population 

growth over time. 

Today 79.8 percent of the population of Latin America live in urban areas, and they account for 12.7 percent 

of the world’s urban population. Forecasts for 2050 indicate a further increase of the share of the urbanized 

population in LAC to 86.2 percent. The degree of urbanization is well above the global average of 54.0 

percent. Figure 12.2 illustrates the extent of urban areas and of urbanization for LAC countries. In relation to 

population, Brazil, Argentina and Mexico have the largest urban areas, while the highest rates of urbanization 

are associated with small states with high population densities. 

1 http://www.un.org/en/development/desa/population/events/ot her/10/index.shtml 


in Latin America and the Caribbean 

373

50 



40 

Urban Area (thousands SQKM) 

30 

20 


10 

0 

Brazil 

Argentina 

Mexico 

Venezuela 

Chile 

Peru 

Colombia 

Ecuador 

Bolivia 

Paraguay 

Uruguay 

Puerto Rico 

Cuba 

Dominican Republic 

Costa Rica 

Panama 

Trinidad and Tobago 

Jamaica 

Guatemala 

Guinea 

Haiti 

Netherlands Antilles 

Nicaragua 

El Salvador 

Honduras 

Guyana 

Suriname 

Aruba 

French Guiana 

Bahamas, The 

Guadeloupe 

Martinique 

Barbados 

Virgin Islands 

Belize 

Antigua and Barbuda 

St. Lucia 

Cayman Islands 

Grenada 

St. Vincent and the Grenadines 

Dominica 

St. Kitts and Nevis 

Anguilla 

Montserrat 

Sao Tome and Principe 

British Virgin Islands 

St. Pierre and Miquelon 

Turks and Caicos Islands 

Figure 12.2 Extent of the urban area and the urbanization index for Latin American and Caribbean countries. 

12.4 | Major threats to soils 

Among the general threats to soil that occur in the region, the three most important ones in LAC will be 

discussed here in more detail. 


Soil erosion by water is the main soil degradation process worldwide and in LAC as well. Wind erosion is also 

prevalent in specific areas with arid and semiarid climates (rainfall lower than 600 mm). 

A high proportion of the land in LAC is on steep slopes, and the main limitation for its agricultural use is 

water erosion (Alegre, Felipe-Morales and La Torre, 1990; Pla, 1993, 1996a; Duvert et al., 2010; PNUMA-CEPAL, 

2010). However, the problems of accelerated water erosion are not confined to steep slopes, but are also 

widespread in agricultural areas with more gentle slopes. In general, the increasing trend of erosion in LAC 

is mainly due to the fast growth of the human population and to pressures put on the land by deforestation 

and over-grazing, and by inappropriate agricultural practices in both subsistence and large-scale high-input 

commercial agriculture (Pla, 1996a). 

Erosion in LAC is mostly caused by water. Some estimates suggest that 42 percent of flood events contribute 

to 70 percent of sediment export (Duvert et al., 2010). In drier areas of Mexico and Argentina on the other 

hand, wind erosion prevails. Estimations of areas affected by erosion in the region vary, but a conservative 

figure is around 15 percent for South America and 26 percent for Central America (Oldeman, 1991b). 

Although there is clear evidence that large and increasing areas of land are being affected by different 

processes of soil erosion, most of the existing evaluations of the type, extent and intensity of soil erosion at 

country or regional level are not very precise or objective. Mass and landslide erosion processes are usually 

not differentiated from surface erosion problems (Hincapié and Ramirez, 2010), leading to an often faulty 

identification of the origin of erosion processes (Pla, 1992, 1993, 2011). The magnitude of the soil erosion 

problems is highly variable, with estimated values for average soil losses in the Andes and Central America 

ranging between 5 and 50 percent of the area or from 100 to 1 000 Mg km -2 yr -1 . Probably almost half of 

agricultural lands are negatively affected by surface soil erosion at different levels, with 15-25 percent, 

depending on the region, strongly affected (Oldeman et al., 1991a, b). 


374 


New regional information has been generated largely through indirect or remote sensing, usually without 

sufficient ground-truthing (Bai et al., 2008; Nachtergaele, Petri and Biancalani, 2011). Some indirect recent 

evaluations based only on soil cover and slope show a reduction in the area with risks of soil erosion to only 5 

percent (EC, 2013). In those evaluations, the processes of mass erosion have not received any attention or have 

been confused with the very different processes of surface erosion (Pla, 2011). 

The control of soil erosion and of the derived effects depend on appropriate land use and management 

planning supported by appropriate soil governance (FAO, 2012). In many LAC countries, the application of 

conservation measures is limited by lack of integration between conservation and development, the lack 

of legislation or ways to implement it, and the shortage of basic local information, trained personnel and 

financial resources (Pla, 1996b). With few isolated exceptions, there have not been policies and well-targeted 

subsidies or incentives through marketing prices and credits to induce sustainable land management. In 

addition, agricultural research has been oriented towards increasing productivity through the use of inputs 

rather than to sustainable land use. This has contributed indirectly to growing environmental problems, 

including soil erosion. 

A main objective of research on soil erosion in LAC must therefore now be to collect and evaluate data to 

generate technology for prediction and control of soil erosion. An understanding of the basic erosion processes 

in each particular situation is required in order to select an effective technology and transfer it to farmers (Pla, 

2003). Some empirical models to predict erosion (USLE, RUSLE) are currently used indiscriminately, without 

scientific evidence of their applicability to a particular situation. The uncritical use of results from these models 

has led to gross errors in planning land use and management (Pla, 2011). However, the RUSLE equation has 

been successfully adopted as a basis for soil use regulations in some countries (Alegre, Felipe-Morales and La 

Torre, 1990). 

One key area for research is the study of existing indigenous practices. An understanding of the biophysical 

and human factors behind these practices might indicate how they can be adopted or adapted to the present 

socio-economic situation. 

Finally, institutional support is essential for developing and assuring continuity of the required research in 

soil erosion and conservation practices in LAC countries (Pla, 2003). 


Soil organic carbon is the main component of humus. It is an organic compound with a stable C/N ratio 

which varies between 10 and 13 in most soils of LAC and of the world as a whole (Palm and Sanchez, 1990; Sisti 

et al., 2004; Jantalia et al., 2007). It is therefore only possible to increase the SOC content if the N content also 

increases to maintain the relationship (Boddey et al., 2012 a, b, 2014). The reverse is also true: increasing one 

unit of soil N availability due to the mineralization of soil organic nitrogen will release between 10 and 13 units 

of C, equivalent to around 40 kg of CO 2 

. 


375 


Figure 12.3 shows soil organic carbon contents and stocks (taking into account soil bulk density) in different Mexican ecosystems. 

Carbon concentrations (left) and carbon stocks (right) in the main ecosystems of Mexico. In both cases the bars with the strongest tone indicate a 

primary forest, closed pasture or permanent agriculture. Bars with the softer tone indicate a secondary forest, open pasture or annual agriculture. Source: 

Cruz-Gaistardo, 2014. 


376 


Primary forests have higher carbon concentration (in percentage) than secondary forests. However, the 

latter are better sinks because they have a greater capacity for conversion of CO 2 

to biomass since they are in 

a more active phase of growth (Vargas et al., 2008). About 185 Pg of organic carbon is stored at 1 m depth in 

LAC soils (16 percent of the world’s soil carbon reserve). Half (95 Pg) of this amount is stored in the soils of the 

Amazon region, and about 52 percent of this carbon pool is held in the top 0.3 m of the soil profile (Batjes and 

Dijkshoorm, 1998). 

LAC biomes where there is high risk of carbon losses – and also biodiversity losses – are the Amazon and the 

Atlantic Forest of Brazil, the Pampas of Argentina, the west coast of Colombia and the core of the Sierra Madre 

del Sur and Sierra Madre Oriental areas in Mexico (Hansen et al., 2013). Satellite images reveal that more than 

half a million square kilometers of Amazon rainforest was destroyed between 1984 and 2005 and replaced by 

agriculture and the introduction of more than 240 million head of cattle (Gardi et al., 2014). 

The richest organic carbon soils in LAC, with stocks higher than 250 tonnes per hectare, are located in the 

sedimentary region of the Carso Huasteco and the Peninsula of Yucatán in Mexico, the tropical forests of 

Guatemala and Costa Rica, the region of Cauca and Magdalena in Colombia, the Orinoco delta in Venezuela, 

in the eastern Amazon, the Uruguayan savannas, the Valdivia forests in Chile, and the wet grasslands and 

steppes of Patagonia in Argentina (Gardi et al., 2014). In these regions the highest proportion of Histosols, 

Andosols and Gleysols with high concentrations of soil carbon is found. The lowest soil carbon stocks are 

found in arid regions of LAC (the deserts of Mexico, Peru and Chile, as well as the drier regions of Brazil and 

Argentina). In these arid regions, it is essential to preserve the scarce carbon (less than 20 tonnes ha -1 ) due to 

the fragility of the ecosystems found there (Figure 12.4). 


377 


Figure 12.4: Organic carbon stock (or density) in soils of Latin America and the Caribbean, expressed in Gigagrams per hectare. 

Source: Gardi et al., 2014. 

Land use changes from forestry to urban or livestock use cause the greatest loss of soil carbon in LAC (Lal, 

2005, 2006; van der Werf et al., 2010; INPE, 2010). The loss of ground cover due to deforestation exposes 

the soil to direct precipitation that could cause erosion and compaction of the soil surface microstructure in 

addition to carbon loss. Deforestation of tropical forests prevents the return to the soil of about 15 tonnes of 

organic inputs per ha eachyear. Agricultural soils return on average only 2 tonnes of residues per ha each year 

(Hughes, Kauffman and Jaramillo, 1999). 

Forest cover loss in the global tropical rainforest biome accounts for about one third (32 percent) of all 

global forest cover loss, and nearly half of this loss of tropical rainforest occurs in South American rainforests. 

Several studies have been carried out and information is documented in the region related to deforestation 

and reforestation (CIAT, 2014; Gardi et al., 2014). The tropical dry forests of South America had the highest rate 

of tropical forest loss, due to deforestation dynamics in the Chaco woodlands of Argentina, Paraguay and 

Bolivia (Grau and Aide, 2008; Viglizzo and Jobbagy, 2010). Brazil is a global exception in terms of forest policy 

change, with a dramatic policy-driven reduction of deforestation in the Amazon Basin (INPE, 2010, 2013; 

Spavorek, 2012) (Figure 12.5). 


378 


Prepared by C. Cruz-Gaistardo 

Figure 12.5: Tree cover in the tonne 2000 and forest loss in the period 2000-2014. (A) Brazil, centered at 5.3°S, 50.2°W; (B) Mexico and 

Guatemala, centered at 16.3°N, 90.8°W and (C) Perú, centered at 8.7°S, 74.9°W; (D) Argentina, centered at 27.0°S, 62.3°W and (E) Chile, 

centered at 72.5°S, 37.4°W. Source: Hansen et al., 2013. 

The litter on the ground also influences carbon fluxes to the mineral soil. A conserved forest can accumulate 

up to 10 tonnes of litter per hectare; however, when this forest is degrading its accumulation rate drops to 2 

tonnes per ha (Cruz-Gaistardo, Díaz and Martínez, 2010). In an extreme case of erosion, litter would completely 

disappear. The loss of soil carbon results in the loss of natural fertility and of the existing biodiversity. These 

factors lead to extreme soil degradation, and could affect the local and regional economy (Smith et al., 2007). 

In the temperate Pampas of Argentina and Uruguay, conversion of grassland and pastures to agriculture 

caused soil organic carbon stocks to decrease from 27 Mg ha -1 to 19-20 Mg ha -1 in the first 20 cm of soil profile 

(Alvarez et al., 2009). 


379 


Across LAC, there are initiatives to improve soil health and the environment. The large cities in LAC, such as 

Mexico City, Lima, Buenos Aires, São Paulo, Rio de Janeiro, Santiago and Bogota, manage their urban wastes, 

which are rich in organic carbon, in order to reduce pollution, especially in the surrounding rural areas which 

typically receive the waste (Da Silva and Donini, 2008). In the region of Santiago de Chile, for example, about 

200 Gg of urban sludge and muds are produced each day by the seven million inhabitants (CEPAL, 2010). These 

wastes are currently being evaluated as possible fertilizers for farmland. However, this type of sludge may need 

up to threeyears to be mineralized, dissolved and available to plants. There is a risk that the concentration of 

metals (mainly zinc and copper) may increase in the soil (González, 2008). In Brazil, there are initiatives to 

grow rice without flooding the soils (so reducing methane emissions) and to cut sugar cane without resorting 

to direct burning. In Colombia and Cuba the use of organic fertilizers is widely promoted (Willer and Lukas, 

2009). 

International policy initiatives such as the United Nations Framework Convention on Climate Change 

(UNCCCD) and the programme for Reducing Emissions from Deforestation and Forest Degradation (REDD) 

often lack the institutional, investment and scientific capacity to begin implementation. In effect, policy is 

sometimes ahead of operational capabilities, even when the necessary information is available. By contrast, 

Brazil’s use of Landsat data in documenting trends in deforestation was crucial to its policy formulation and 

implementation. To date, only Brazil produces and shares spatially explicit information on annual forest extent 

and change (UNFCCC, 2013; INPE, 2010, 2013). 

Most soil sampling sites in LAC have been selected to address questions related to fertility or taxonomy. 

They were not selected to specifically quantify carbon in different pools (above- and below-ground). 

Consequently, they cannot be used for payment for environmental services. To overcome this, Mexico and 

Brazil are currently conducting national soil surveys designed to quantify soil carbon, increasing the number 

of sites within a systematic grid and using state-of-the-art instrumentation. The surveys are employing 

field spectrophotometers and micrometeorological towers to measure carbon fluxes between the land and 

atmosphere (Vargas et al., 2013). In the case of Mexico, there are currently 11 000 sites with systematic 

information on all carbon reservoirs and coordinated programs for re-sampling and continuous monitoring. 

12.4.3 | Soil salinization 

Salt-affected soils are found mainly in the arid and semi-arid regions. Salinization caused by irrigation 

affects 18.4 million ha in LAC, particularly in Argentina, Brazil, Chile, Mexico and Peru (AQUASTAT, 1997). Many 

of the large plains of the continent are affected by natural salinization, but land use changes and overgrazing 

have also caused topsoil salinization (Taboada et al., 2011; Bandera, 2013; Di Bella et al., 2015). 

Around 85 million ha are affected by excess of salts and sodium in Argentina, including arid and semi-arid 

zones (Szabolcs, 1979). Information provided by Bandera (2013) indicates that Argentina has approximately 

600 000 ha of irrigated soils affected by salinity, which is the third largest area in a single country after Russia 

and Australia. In Argentina there are also areas of soils affected by salts in humid and sub humid climates, 

where salts come from groundwater. This is the case of the Pampa Deprimida (Depressed), the Buenos Aires 

East-Center (9 million ha), Buenos Aires Northwest (2.5 million ha), and the Submeridional Lowlands of Chaco 

and Santa Fe provinces (3 million ha). 

The irrigated agricultural area of the northeast region of Brazil is around 500 thousand ha, and 25–30 

percent is in the process of salinization (Heinze, 2002). Irrigation-induced salinization is an important land 

degradation process that affects crop yield in the Brazilian semi-arid region. Gypsum has been used as a 

corrective measure for saline soils in these areas (Moreira et al., 2014). The addition of gypsum to irrigation water 

improves soil physical and chemical properties and can be considered as an alternative for the reclamation of 

saline-sodic and sodic alluvial soils in Northeast Brazil (Silveira et al., 2008). 


380 


The salinity of soils on the cultivated land in the coastal valleys in northern Chile (Azapa and Lluta valleys) is 

generated by the salinity of irrigation water. 

In Colombia, soils susceptible to salinization cover approximately 86 592 km 2 , of which 90 percent are 

located in dry regions (Casierra-Posada, Pachón and Niño-Medina, 2007). Areas susceptible to salinization are 

located in the Caribbean region and inter-Andean valleys and highlands. 

In Cuba, the soils with primary salinity only occupy restricted areas, often near to sea coasts. Secondary 

salinity, however, affects the majority of Cuban soils. The main causes of secondary salinity are: increase of the 

level of saline groundwater, deforestation of hilly lands, and use of saline water for irrigation (Alvarez et al., 

2008). There are 160 000 ha under rice, of which approximately 100 000 ha have problems of salinity and/or 

sodicity, in varying degrees (Borroto and Castillo, 1986). 

In Ecuador, approximately 8.1 million ha have soils suitable for agriculture, of which about a quarter are 

currently used for agriculture. This includes: production of short cycle crops such as rice, maize, cassava, soy, 

watermelon, melon, tomato, pepper; cultivation of perennial crops, including cocoa, coffee, sugarcane and 

banana; and, on 63 percent of the area, natural and artificial pastures. One of the most serious problems 

facing this ecosystem is the increase in salinity. This is caused by irrigation with waters of medium quality 

and by excessive use of fertilizers in fert-irrigation. Soil salinity constitutes one of the main causes of crop 

yield reduction, and a significant part of Ecuadorian highland soils are considered as having high-salinity. The 

main causes are the pyroclastic nature of the soils, the effects of erosion, and the poor use of irrigation water 

(Jaramillo, Arahana and Torres, 2014). 

In Mexico, approximately 20 percent of irrigated agricultural land (6 million ha) is affected by salinity and 

sodicity problems. In 1964 it was estimated that, in the irrigation districts of Culiacán, Fuerte River, Río Mayo, 

and Rio Yaqui which have a total area of 610 701 ha, over one third (218 495 ha, 36 percent) of the area had 

some kind of salt problem (Palacios-Vélez, 2012). 

The lack of complete feasibility studies prior to implementing irrigation projects in Peru has caused 

increased drainage and salinity problems on 2 500 km 2 in the Coastal Area. In fact, all irrigation projects on the 

coast have experienced drainage and salinity problems within a fewyears after starting to irrigate (Cornejo, 

1970). More than 300 000 ha (40 percent of the cultivated land in Peru’s Coastal Valleys) were affected by soil 

salinity in the 1970s. About 25 percent (roughly 190 000 ha) were characterized by light to extreme salinity 

(>4 dS m -1 ), enough to have negative effects on crop productivity (World Bank, 2007). Currently, there is little 

interest from the government to stop degradation of land. Even up-to-date information is lacking and what 

there is dates back to the 1970s (Santayana, 2012). 

In Venezuela, the existence of problems of salinity was recognized in soils in the arid and semi-arid areas 

of the States of Zulia, Falcón, Lara, Anzoátegui and Trujillo (Villafañe, 1995). Falcon was one of the major 

vegetable producing states in Venezuela until soil degradation by salts had negative effects on yields and 

caused the abandonment of farms in this area. 


381 



12.5.1 | Argentina 

Argentina is the eighth largest country in the world, with an area of 2 780 400 km 2 . Seventy percent of 

its territory has an arid, semiarid or dry sub-humid climate, and the remaining 30 percent has a humid or 

sub humid climate. Regions under humid and sub humid temperate and subtropical climates (e.g. Pampean, 

Chaco and Mesopotamia) concentrate on the production of cereals, oilseeds, industrial crops, forages, forest 

plantations, domestic livestock and dairy products (SIIA, 2015). 

Agriculture in Argentina began in earnest at the end of 19th century with the arrival of European immigrants 

and government colonization policies (Barski and Gelman, 2001; Viglizzo and Jobbagy, 2010). Agriculture 

and livestock grazing expanded until the mid-20th century by bringing new lands into production, largely 

employing low intensity production practices (Viglizzo and Jobbagy, 2010). This resulted in moderate to severe 

land degradation, not only in agricultural areas but also in dryland areas (SAGyP-CFA, 1995). In the second half 

of the 20th century agriculture intensified, especially in the Pampean region, with the use of more productive 

cereal varieties, hybrids and genetically modified crops, fertilizers and no till farming (Paruelo, Guerschman 

and Verón, 2005; Satorre, 2005; Viglizzo and Jobbagy, 2010). 

In recentyears, the agricultural frontier has expanded to the north-east, the north-west and the west 

(Figure 12.6), moving into areas with drier climates and/or less fertile soils (Paruelo, Guerschman and Verón, 

2005; Viglizzo and Jobbagy, 2010). As a result, the cultivated area increased from about 15 to 32 million ha from 

1988 to 2010, and bulk grain production shot up from about 20 to nearly 100 million tonnes in the same period 

(SIIA, 2015). At the same time, the ratio of crops produced changed. In 1990, the mix was: 37 percent wheat, 30 

percent soybean and 13 percent maize. Twenty fouryears later (in 2014), production was 61 percent soybean, 19 

percent maize and only 11 percent wheat (SIIA, 2015). This shift was driven by export demand for oil and biofuel 

soybean (Gobierno Argentino, 2014). Although a success in terms of fuel saving and adoption by farmers, this 

move towards a soybean monoculture appears to have driven many smaller farmers out of business (Pengue, 

2005). 


382 


Figure 12.6 Expansion of the agricultural frontier under rainfed conditions in the north of Argentina. Source: Viglizzo & Jobbagy, 2010. 

It has been estimated that agriculture in Argentina during the 20th century decreased soil carbon stocks by 

27-50 percent. The causes included the turning over of grasslands and prairie soils rich in organic C with mould 

board and disc ploughs, and the loss of crop diversity (Studdert, Echeverría and Casanovas, 1997; Viglizzo and 

Jobbagy, 2010; Sainz Rozas, Echeverria and Angelini, 2011; Caride, Piñeiro and Paruelo, 2012; Milesi Delaye et 

al., 2013). The potential of no-till farming to increase soil organic C stocks is still under discussion. In a review 

of 42 no-till vs plough till data sets from field experiments conducted in the Pampean region, Steinbach and 

Álvarez (2006) concluded that a 2.76 Mg ha -1 organic C increase was observed in no-till systems compared with 

tilled systems, with a relatively higher increase of organic C in areas of low organic matter level. However, 

Álvarez et al. (2009) observed no significant change in topsoil organic C content after adoption of continuous 

no-till in previously conventionally tilled loam, silty loam, and silty clay loam soils in well-managed farms of 

the region. 

No-till farming is considered to improve topsoil physical properties, especially when combined with suitable 

crop rotations and pastures (Álvarez et al., 2014). However, the prevalence of soybean mono-cropping in the 

Pampean and Chaco regions promoted unfavourable topsoil physical conditions such as laminar and massive 

aggregation, shallow compaction and decreased infiltration rates (Sasal, Andriulo and Taboada, 2006; Álvarez 

and Steinbach, 2009; Álvarez et al., 2009, 2014). These structural forms were found to decrease soybean yields 

under rainfed conditions (Bacigaluppo et al., 2011). 


383 


Fertilizer use in Argentina was quite low until the mid-1990s and despite the eight-fold increase in fertilizer 

use in the last two decades, negative budgets of nitrogen, phosphorus and sulphur still persist in the Pampean 

region (Lavado and Taboada, 2009; Viglizzo and Jobbagy, 2010). Nutrient removal by grains was found to have 

exceeded application by 2.3-3.2, 1.4-2.0 and 2.0-2.6 times for N, P and S over a fouryear period, respectively. 

Phosphorus is a special issue, as it has decreased to ‘deficient’ levels in several areas (Sainz Rozas, Echeverria 

and Angelini, 2012). 

During the 20th century, more than 1.2 million ha (32 percent of the agricultural area) were affected by 

moderate to severe soil water erosion, characterized by 47 to 131 Mg ha -1 yr -1 soil losses and by 5 to 20 cm surface 

horizon thickness decreases (SAGyP-CFA, 1995). According to estimations by Viglizzo and Frank (2010), the 

widespread adoption of no-tillage helped to control erosion losses, which was reported to have decreased to 

only 7 Mg ha -1 yr -1 . However, more field data are needed to check these estimations. 

Salt affected soils cover more than 70 million ha in Argentina, the third largest area in any country in the 

world (FAO, 1976). At least 600 000 ha of irrigated soils under arid and semiarid climates are affected by 

anthropogenic salinization, often related to unsuitable drainage management and/or poor water quality. 

In sub humid and humid regions, there are about 12 million ha of naturally salinized soils. In these areas, 

anthropogenic salinization was also promoted by livestock grazing (Taboada, Rubio and Chaneton, 2011), 

supplementary irrigation (Andriulo et al., 1998) deforestation (Nosetto et al., 2012) or afforestation (Jobbagy 

and Jackson, 2004). 

A detailed evaluation of the state of land degradation in the drier areas of Argentina came to the following 

conclusions (FAO, 2010): 

• Eighty one percent, or more than 1 240 000 km 2 of the evaluated systems, present some process of 

degradation. The degree of degradation of dry lands, defined by the intensity of the process, varies from 

‘non-degraded’ to ‘extreme’. This result is obtained after adding all degradation processes identified, 

including water and wind erosion, salinization and loss of vegetative cover. 

• Amongst the degrading processes analyzed, 50 percent of the area is affected largely by processes 

of biological degradation (variation of vegetative cover, loss of habitats, and decrease of biomass). 

Twenty 6 percent of the area presents a strong degree of degradation. The rate of degradation is slow 

in 40 percent of the areas with biological degradation, while almost 20 percent of these areas present 

a high rate of increase. 

• Fifteen percent of the land use systems analyzed presented symptoms of physical degradation of the 

soil (compaction, crusting, flooding). Of these 60 percent present a high degree of degradation. Thirty 

percent are degrading at a moderate rate. 

• Five percent of the area analyzed presented symptoms of degradation of water resources (decrease 

of the average moisture content of the soil, changes in volume, reduction of the quality of water, 

reduction of the capacity to capture and retain water). Eighty percent of the area presents a strong 

degree of degradation, with approximately 20 percent having a moderate rate of increase. 

• Forty percent of the area analyzed is affected by wind erosion, with a degree of degradation that is 

mostly moderate to strong with a tendency to increase. Water erosion affects 40 percent of the area 

analyzed (Figure 12.7), with a similar degree of severity to wind erosion, although with a lower rate of 

increase. 


384 


Figure 12.7 Percentage of areas affected by wind (a) and water erosion (b) in Argentina. Source: Prego et al., 1988. 


385 


12.5.2 | Cuba 

The main island of Cuba is the largest island in the West Indies, with a total land area of 104 945 km². 

The area of the country as a whole, including the Isla de la Juventud (2 200 km 2 ) and around 4 195 keys and 

small islands, is 110 860 km 2 . The topography is mostly flat to rolling, with rugged hills and mountains in 

the southeast and south-central area. The Cuban mountain range system is formed by four massifs covering 

about 18 percent of the surface of the Cuban archipelago. The surface cover of Cuba consists of cropland and 

crop/natural vegetation mosaics (44 percent), shrub lands, savanna and grasslands (24 percent), forests (23 

percent) and wetlands (9 percent). 

Mean annual rainfall is 1 335 mm, with a pronounced seasonal variation between the driest and wettest 

months. Rainfall levels vary widely across the country, from 300 mm annually in the Guantánamo area of the 

south to more than 3 000 mm in the north. Mean annual temperature is 25°C (CUBA, 2014). 

In recent decades, significant variations have been detected in the country’s climatic patterns (Centella et 

al., 1997). An overall increase in temperature has been accompanied by a reduction in annual rainfall totals of 

10-20 percent and an increase in inter-annual variation in rainfall of 5-10 percent, with reduced rainfall in the 

rainy season and increased rainfall in the dry season (Lapinel, Rivero and Cutié, 1993; Goldenberg et al., 2001). 

At the same time, the frequency of unseasonal droughts has increased. 

The National Environment Strategy 2007/2010 is the guiding document for Cuban environmental policy. 

It defines the five main environmental issues in Cuba, which are: land degradation; factors affecting forest 

coverage; pollution; loss of biological diversity; and water scarcity. The Strategy proposes policies and 

instruments to tackle the five issues in order to improve environmental protection and the rational use of 

national resources. Land degradation is considered the most important of the five issues. 

Cuba has a wide variety of soils, including those developed over sedimentary limestone (Nitisols, Ferralsols 

and Lixisols), and others over older rocks (Cambisols and Phaeozems). Cuba has complete soil studies and 

maps at cartographic scales of 1:250 000, 1:50 000 and 1:25 000 (Gardi et al., 2014). 

The most recent assessment of land degradation at local and national level was carried out between 2006 

and 2010, using a standard methodology (Liniger et al., 2011; Bunning, McDonagh and Rioux, 2011). In parallel, 

the sustainable land management practices applied to stop or reverse the degradation process in the country 

were inventoried in each land use system (AMA, 2010). A first set of 23 thematic maps at 1: 250 000 scale was 

produced in 2010 and these were updated in 2013 (IGT-AMA, 2014). 

Fifteen different types of degradation, affecting the whole country to a lesser or more serious degree 

were recognized (Figures 12.8 and 12.9). The four main types of degradation in terms of extent are: (1) loss of 

topsoil by water erosion, which covers more than 30 000 km 2 and is present in all provinces; (2) the loss of 

vegetative cover, which has a similar extent and also occurs in every province; (3) processes of salinization and 

compaction, which cover between 10 000 to 20 000 thousand km 2 in 14 and 11 provinces respectively; and 

(4) loss of habitat condition and areas affected by fire, which occupy up to 5 000 km 2 each in nine provinces of 

the country. The other types of degradation such as aridity, loss of soil fertility, reduced organic matter content 

and reduced quality of surface and groundwater occupy areas that do not exceed 5 000 km 2 . 


386 


Figure 12.8 Predominant types of land degradation in Cuba. Source: FAO, 2010. 

Figure 12.9 Extent of land degradation in land use system units in Cuba. Source: FAO, 2010. 

As far as the intensity of degradation is concerned, 12 percent of Cuban territory is classified as grade 1 

(slight), while 68 percent is grade 2 (slight or moderate intensity) and 19 percent is grade 3 (strong or intense 

degradation). The last two grades are mainly located in the central and eastern area (Figure 12.10). 


387 


Figure 12.10: Intensity of land degradation in Cuba. Source: FAO, 2010. 

A systematic local inventory of land degradation processes was carried out in the dryland areas of Cuba. 

The drylands occupy a total area of nearly 10 000 km 2 , and occur mainly in the eastern region of the country, 

between Camagüey and Guantanamo, covering 6 provinces and 20 municipalities. Surveys were carried 

out in the most representative areas of Camaguey-Tunas, Granma and Guantanamo covering soil health, 

water quality and quantity, vegetation status and biodiversity, among others. Since 2001, the country has 

established different programmes and strategies based on sustainable soil management approaches in order 

to combat soil degradation. The most important of these is the National Program for Soil Improvement and 

Conservation, which is overseen by the Soil Institute. Over the last decade, at least 500 000 ha have benefited 

from these programs, which are financially supported by Cuban government (Instituto de Suelos, 2001). 

12.6 | Conclusions and recommendations 

LAC has a wide range of biomes and probably the largest potential area of cropland in the world. The soils 

of the region are subject to a number of threats, of which the three most important are: soil erosion, organic 

carbon losses and salinization. Other threats such as imbalance of nutrients, loss of biodiversity, compaction, 

waterlogging, contamination, and sealing and capping are also common (Table 12.1). 

The most important ecosystem services affected in LAC are: climate regulation through the disturbance 

of the C and N cycles due to deforestation, and water regulation and food production on sloping lands. 

Deforestation and erosion caused by inappropriate land use are the initial causes of various anthropogenic 

threats to soil quality. 

An enhanced natural resource information system is necessary in many countries of LAC, in order to perform 

a better diagnosis of soil conditions and their level of degradation. This will allow possible solutions to be 

identified, including land use planning and appropriate legislation. 

A major effort is required to design and implement sustainable soil management in the region, taking into 

account the risks and threats assessed as well as the particular characteristics of each country. A participatory 

process is required if the final goal is to be reached, namely, protection of the soil resource for food security and 

for the production of ecosystem services for present and future human well-being. 


388 


Threat to 

soil function 


Organic 

carbon change 

Salinization 


Nutrient 

imbalance 

Loss of soil 

biodiversity 

Compaction 

Waterlogging 


acidification 

Contamination 


and land take 

Summary 

Widespread across the 

region. 

Landslides are accelerated 

by land use in highland 

areas 

Declines are caused by 

deforestation, 

intensive cultivation 

of grasslands and 

monoculture. 

Caused by inadequate 

irrigation technology 

and water quality. Land 

use changes also promote 

salinization. 

Most countries have 

negative nutrient balances 

due to over-extraction. 

In some cases over 

fertilization also causes 

nutrient imbalance. 

Suspected to occur in 

deforested and overexploited 

agricultural 

areas. 

Caused by overgrazing 

and intensive agricultural 

traffic. 

Due to deforestation and 

poor structural conditions 

in agricultural areas. 

Soil acidification is limited 

to some areas with 

overuse of N fertilizers 

Industrial sources cause 

soil contamination in 

some places. 

Non-point soil pollution 

prevails in sites with 

intensive agriculture 

(e.g. herbicides residues). 

In some valleys and 

floodplains, urbanization 

has expanded onto fertile 

soils. 


Very poor Poor Fair Good Very good 

Confidence 

In 

In trend 

condition 

Table 12.1 Summary of Soil Threats Status, trends and uncertainties in in Latin America and the Caribbean 


389 


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of Soil Changes in the Near East 

and North Africa 

Regional Coordinator: Seyed Kazem Alavi Panah (ITPS/Iran) 

Regional Lead Author: Mubarak Abdelrahman Abdalla (Sudan) 

Contributing Authors: Hedi Hamrouni (Tunisia), Abdullah AlShankiti (ITPS/Saudi Arabia), Elsiddig Ahmed 

El Mustafa El Sheikh (ITPS/Sudan), Ali Akbar Noroozi (Iran). 


399 

in the Near East and North Africa


The Near East and North Africa (NENA) region includes Tunisia, Algeria, Morocco, Libya, Egypt, Sudan, 

South Sudan, Jordan, Israel, Lebanon, Syria, Palestine, Iraq, Yemen, Saudi Arabia, Oman, Qatar, United Arab 

Emirates, Kuwait, Bahrain and Iran. The region has a land area of approximately 14.9 million km 2 , nearly all of 

which is hyper-arid, arid or semi-arid. The region faces three climatic constraints: aridity, recurrent drought, 

and desertification, the latter also in part human induced. South Sudan is the only country in the region which 

falls within the dry sub-humid tropical zone. Large areas of Libya, Egypt, Bahrain, Kuwait, Qatar and the 

United Arab Emirates are entirely desert (FAO, 2013a). 

The soils of the region are broadly as follows: 

• The soils of the Maghreb region (Morocco, Tunisia, Libya and Algeria) fall into three broad divisions: (i) 

along the Mediterranean and Atlantic coasts productive Kastanozems (Xerolls) and Luvisols (Alfisols) 

occur (SEDENOT, 1999; MADRPM, 2000; Halitima, 1988); (ii) Leptosols (lithic subgroups) and Cambisols 

(Inceptisols) are found in the Atlas Mountains away from the coast (Yigini, Panagos and Montanarella, 

2013); and (iii) Calcisols (Calcids), Gypsisols (Gypsids), Leptosols and Cambisols are found in the southern 

part (Jones et al., 2013). 

• Vertisols, Arenosols (Psamments), Fluvisols (Fluvents), Calcisols and Gypsisols (Aridisols) are the 

dominants soils in Sudan and Egypt. 

• In the Mashreq region (Jordan, Syria, Lebanon, Iraq and Palestine), the soils of the valleys are Arenosols 

(Psamments) and Fluvisols (Fluvents). In the highlands, steppe and desert regions, the main orders are 

Calcisols (Calcids) and Cambisols (Aridisols), Arenosols (Psamments) and Leptosols (Lithic subgroups), 

and Vertisols which are calcareous in the subsoil horizons. 

• In the Arabian Peninsula and the Gulf (Oman, Kingdom of Saudi Arabia, Kuwait, Bahrain, United Arab 

Emirates, Yemen, Iran and Qatar), there are alluvial soils rich in silt and desert soils, and sandy soils 

poor in organic carbon but in which evaporite Tertiary Formations played an important role in the 

formation of contemporary minerals (Abbaslou et al., 2013). 

Agriculture is an important source of income for many countries in the region. Arable land constitutes only 

6.8 percent of the total land area, while about 26 percent is used for pasture and about 7 percent is under forest 

(Hamdallah, 1997). Based on the type of agriculture practiced, Dregne and Chou (1992) divided the productive 

land in the region into: irrigated land (0.7 percent, 8.5 million ha); rainfed (2.2 percent, 27.7 million ha); 

rangelands (40 percent, 495.6 million ha); and extremely arid land (57 percent, 705.2 million ha). Proportions 

vary considerably by country. Productive lands represent 30 percent of the total area in Syria and Lebanon, 

but only 3 percent in Egypt, Algeria and Sudan, and only 0.5 percent in Saudi Arabia, Oman and Mauritania 

(Mamdouh Nasr, 1999). 

Natural resource degradation, especially where agriculture is practiced, is a real threat in all countries of the 

region and remains a major limitation to the reliable supply of food. In most countries, salinization, water and 

wind erosion, loss of vegetation cover, soil physical degradation (including compaction and surface crusting) 

are the main threats to the soil’s capacity to provide ecosystem services. The expansion of agriculture into 

marginal lands has greatly aggravated water erosion and consequently soil degradation. In almost all countries 

in the region, extreme climatic conditions, overgrazing, unsuitable cropping patterns and accumulation of 

salts have rendered large areas of land unproductive (Abahussain et al., 2002). 

Mamdouh Nasr (1999) reported that rainfed cropland represents only 3 percent (about 30 million ha) of the region’s 

total drylands, yet about 22 million ha of this total (73 percent of the cropland area) are estimated by UNDCPAC to be 

degraded. The extent of degradation of rainfed cropland is greatest in the countries of northern Africa: Algeria (93 

percent), Morocco (69 percent), Tunisia (69 percent) and, exceptionally, Egypt (10 percent). The eastern sub-region 

(Iran, Iraq, Jordan, Lebanon, Sudan and Syria) is the part of the NENA region most affected by land degradation (FAO, 

2004). The extent of degradation in the countries of the Middle East - Iraq (72 percent), Syria (70 percent) - is higher 

than that in the Gulf Countries: Oman (50 percent), Qatar (25 percent), and Bahrain (20 percent). 


400 


The Arab Centre for the Study of Arid Zones and Dry Lands (CAMRE/UNEP/ACSAD, 1996) has estimated that, 

overall, land degradation affects approximately 49 percent of farmland in the eastern sub-region; 29 percent 

in the Nile Valley of Egypt; 17 percent in North Africa; and 9 percent in the Gulf Cooperation Council Countries. 

More than one process of degradation can occur in a single farming system. For example, degradation is 

serious in one of the largest countries in the region, Sudan, a country with high agricultural potential (Ayoub, 

1998). In Sudan, the 46 million ha lying in the semi-arid zone, where mixed farming of both animal husbandry 

and rainfed arable cropping are practiced, have experienced intensive soil degradation over the last 35years, 

affecting production of field crops, gum Arabic, and livestock products. 

In North Africa, causes of soil degradation are divided between overgrazing (68 percent), over-cultivation 

(21 percent), deforestation (10.5 percent) and overexploitation of natural vegetation for about 0.5 percent 

(Thomas and Middelton, 1994). In the newly created country of the region, South Sudan, there are very limited 

studies on land degradation. However, there are indications that land use changes have impaired the quality 

of the land in many places (Dima, 2006). 

Land degradation in certain areas may affect adjacent areas. For example, degradation in rangelands where 

rainfall is low has negative effects on resources of rainfed farming areas. This is also one of the regions most 

vulnerable to climate change (FAO, 2011). Agriculture faces major losses due to land degradation, and yields 

are expected to decrease by the year 2050: rice yields by 11 percent; soybean yields by 28 percent; maize yields 

by 19 percent; and barley grain yields by 20 percent (FAO, 1994). Recent studies on the economic cost of land 

degradation in the region were reported by Hussein et al. (2008); they were estimated at US$9 billion yr -1 (2.1- 

7.4 percent of GDP). 

This chapter will discuss the main soil threats - erosion by water and wind, salinity/sodicity, soil 

contamination, and organic C depletion. Major causes of soil degradation in the region are due to many 

factors, including: (i) excessive irrigation and poor drainage; (ii) wind and water erosion; (iii) waterlogging; 

(iv) deteriorated soil fertility; (v) over-grazing; (vi) loss of soil cover; (vii) land mis-management; (viii) sand 

encroachment; and (ix) overuse of herbicides, pesticides and chemical fertilizers (FAO, 2004). 

Data correlating land degradation with yields are scarce at the global level. However, recent studies (2000- 

2010) show that land losses due to degradation in North Africa are the highest among selected countries 

worldwide, and that this has resulted in a considerable food gap of 0.6×106 metric tonnes (Wiebe, 2003). 

There have been many efforts to tackle the issue of land degradation. Experience has shown that a key factor 

for success is political will. In this respect, the 21 st summit of African leaders in 2013 urged member states to 

place land degradation at the centre of the debate on the post-2015 development agenda, and to recognize it 

as one of the sustainable development goals. This is particularly important for NENA because the agricultural 

sector in the region contributes about 10 percent of the region’s GDP, but is characterized by an exceptionally 

fragile and vulnerable resource base. 


401 


13.2 | Major land use systems in the Near East and North Africa 

Land use systems in the region have been identified and broadly delimited based on a range of characteristics 

(Nachtergaele and Petri, 2011). Their geographical location is indicated in Figure 13.1. They can be combined 

in three major systems: irrigated crop based, rainfed mixed and livestock based. These systems are briefly 

discussed below (Dixon and Gulliver, 2001). 

Irrigated land use systems 

Given the arid and semiarid nature of much of North Africa and the Near East, irrigated farming has always 

been of crucial importance in generating much of the region's agricultural output. The ‘irrigated farming 

system’ in NENA contains both large and small-scale irrigation schemes with high population densities and 

generally very small farm sizes. The prevalence of poverty within both large and small segments of the system 

is moderate. 

Traditionally, areas within the large scale irrigation sub-system have been linked primarily to perennial 

surface water resources, such as the Nile (Egypt and Sudan) and Euphrates (Syria and Iraq). However, the 

intensification of traditional karez or qanat systems has also led to the evolution of large-scale irrigated 

areas where sub-surface water is abundant. More recently, the availability of deep drilling and pumping 

technologies has permitted the development of new areas drawing entirely on subterranean aquifers. Largescale 

schemes are found across all zones of the region and include high-value cash and export cropping and 

intensive vegetable and fruit cropping. 

Patterns of water use vary greatly but throughout the region inappropriate policies on water pricing and 

centralised management systems have meant that water is seldom used efficiently. Significant economic and 

environmental externalities have arisen through excessive utilisation of non-recharged aquifers while, in a 

number of cases, the excessive application of irrigation water has resulted in rising groundwater tables, soil 

salinization and sodification problems. 

The small scale irrigated sub-system also occurs widely across the region. Although not as important as the 

larger schemes in terms of numbers of people involved or in the amount of food and other crops produced, it 

is a significant element in the survival of many people in arid and remote mountain areas. This sub-system, 

examples of which are sometimes of considerable antiquity, typically develops along small perennial streams 

and at oases, or where flood and spate irrigation is feasible. It sometimes also draws on shallow aquifers 

and boreholes, although these rarely penetrate to the depths seen in the large schemes. The major crops 

grown within small-scale irrigation areas are mixed cereals, fodder and vegetables. These areas also provide 

important focal points for socio-economic activity, but intense local competition for limited water resources 

between small rural farmers and other users is becoming increasingly evident. 

Rainfed mixed crop and livestock land use systems 

The rainfed agricultural and livestock systems are the most important land use system in the region in 

terms of population engaged in agriculture. However, as these systems are practiced on less than 10 percent 

of the land area, population densities in these farming areas are moderately high. These systems covers two 

sometimes overlapping segments. The first segment occurs on high terraces and is dominated by rainfed 

cereal and legume cropping, with tree crops, fruits and olives on terraces, together with vines. In Yemen, 

higher reaches are reserved for qat trees and coffee, which are traditionally the most important tree crops in 

Yemen’s mountain regions. The second segment is based primarily on the raising of livestock (mostly sheep) 

on communally managed lands. In some cases, both the livestock and the people who control them are 

transhumant, migrating seasonally between lowland steppes in the more humid winter season and upland 

areas in the dry season. This type of livestock keeping is still important in Iran and Morocco. 


402 


Poverty within this system is extensive, as markets are often distant, infrastructure is poorly developed 

and the degradation of natural resources is a serious problem. In the lowlands where rainfed production is 

feasible, an increasing area is now benefiting from the availability of new drilling and pumping technologies, 

which have made it possible to use supplementary winter irrigation on wheat and full irrigation on summer 

cash crops. There is some dry season grazing of sheep migrating from the steppe areas. 

The more humid areas (with 600 to 1000 mm annual rainfall) that occur in the Caspian and Mediterranean 

coastal areas are characterised by tree crops (olives and fruit), melons and grapes. There is also some protected 

cropping with supplementary irrigation for potatoes, vegetables and flowers. Common crops are wheat, 

barley, chickpeas, lentils and fodder crops. Poverty in these more humid areas is moderate, but would be 

higher without extensive off-farm income from seasonal labour migration. 

Figure 13.1 Land use systems in the Near East and North Africa. Source: FAO, 2010. 

Livestock-based land use systems in sparsely vegetated areasù 

The pastoral land use system, mainly involving sheep and goats but also with some cattle and camels, is 

practiced on large areas of semiarid steppe lands, and is characterised by low population densities, with more 

densely populated areas around irrigated settlements. There are irrigated croplands scattered throughout the 

system, thus boosting the agricultural population – and helping to support a cattle population. Strong linkages 

exist to other farming systems through the movement of stock, both through seasonal grazing of herds in 

more humid areas and through the sale of animals to large feedlots located around urban areas. Seasonal 

migration, which is particularly important as a risk minimisation measure, depends on the availability of 

grass, water and crop residues in neighbouring arable systems. Nowadays, pastoral herds are often partially 

controlled and financed by urban capital. Where water is available, small areas of crop production have been 

developed to supplement the diets and income of pastoral families. However, such sites are few and poverty 

within the system is extensive. 

The sparse (arid) land use system covers more than 60 percent of the region and includes vast desert zones. 

People are concentrated in oases and on a number of irrigation schemes (notably in Tunisia, Algeria, Morocco 

and Libya). Part of the land is irrigated and utilised for the production of dates, other palms, fodder and 

vegetables. Pastoralists within this system also raise camels, sheep and goats. Following scattered storms 

and in good seasons, the system provides opportunistic grazing for the herds of pastoralists. The boundary 

between pastoral grazing and sparse agriculture systems is indistinct and depends on climatic conditions. 

Poverty within this system is generally low as population pressure is limited. 


403 


13.3 | Major threats to soils in the region 

13.3.1 | Erosion 

Water erosion 

Water erosion is predominant in the part of the region which has sloping lands and where rainfed agriculture 

is practiced, although it may also occur in gently sloping areas. The degree of water erosion depends on the 

intensity and duration of the rainstorms, often enhanced by the terrain attributes and land use practices, 

particularly where these have reduced land cover. Water erosion results in the removal of fertile soil and in the 

reduction of irrigation efficiency and storage capacity. Considerable volumes of soil may be lost. Based on the 

GLASOD survey quoted by Abahussain et al. (2002), the total area affected by water erosion in NENA has been 

estimated at about 41 million ha. However, the extent varies significantly by country (Table 13.1.). 

Country Area Country Area Country Area 

Algeria 3 900 Lebanon 65 Sudan 17 300 

Bahrain 0 Libya 1 300 South Sudan n.d. 

Egypt Negligible Morocco 3 600 Syria 1 200 

Iran 1 70 000 Oman 2 800 Tunisia 3 800 

Iraq 1 150 Palestine n.d. 

United Arab 

Emirates 

0 

Jordan 330 Qatar 0 Yemen 5 600 

Kuwait 0 Saudi Arabia 200 

Table 13.1 Land degradation caused by water erosion in the NENA region (1000 ha) 

Source: Abahussain et al., 2002. 

1 Azimzadeh et al., 2008 


404 


Wind erosion 

A review conducted by Abahussain et al. (2002) indicated that more than half of the total area in the 

region received annual rainfall of less than 150 mm. Consequently, large areas are without plant cover, or the 

cover is very sparse. This situation is aggravated further by high land-use pressure, both from human and 

animals, which ultimately causes severe topsoil disturbance. The resulting wind erosion is the most common 

environmental problem in the region and accounts for approximately 60 percent (135 million ha) of soil 

degradation. Countries differ in the extent they are affected, with Saudi Arabia the most affected (Table 13.2). 

Wind erosion has resulted in detrimental effects on land quality by removing the fertile top soils. In addition, 

the accumulation of eroded materials in irrigation canals, agricultural fields (sand encroachment) and water 

harvesting points affects the cropped areas in the region severely. 

Country Area Country Area Country Area 

Algeria 12 000 Lebanon Sudan 71 000 

Bahrain n.d. Libya 24 000 South Sudan n.d. 

Egypt 1 400 Morocco 600 Syria 3 000 

Iran 1 20 000 Oman 4 000 Tunisia 4 000 

Iraq 3 000 Palestine n.d. 

United Arab 

Emirates 

1 100 

Jordan 3 000 Qatar 200 Yemen 6 000 

Kuwait 300 Saudi Arabia 50 000 

Table 13.2 Soil degradation caused by wind erosion in the NENA region (1000 ha). 

Source: Abahussain et al., (2002 ). 

1 Azimzadeh et al., 2008 

An example of the problems is provided by the largest irrigated scheme in the world – the Gezira scheme 

in Sudan – which has been badly affected by sand encroaching from surrounding areas. One study found that 

over the last decade, wind erosion has decreased the soil level outside the scheme by 10 cm, whereas soil 

depth increased inside the scheme by 30 cm - any increase in level beyond 20 cm prevents irrigation water 

flow. Inside the scheme, the topsoil texture has changed from clayey to more sandy. Dune displacement in 

the scheme was about 3-5 meters month -1 during the season of active sand movement (Al-Amin, 1999). This 

situation has forced famers to take considerable land out of production because of problems irrigating, with 

sand filling the irrigation canals and altering slopes (Mohammed, Stigter and Adam, 1995). 

The continental sands of Kordofan and Darfur (Sudan) which support the world’s major source of gum Arabic 

(Acacia senegal L.) encounter severe wind erosion due to overgrazing and mismanagement, which endangers 

the dominance of the species in the area. Omar et al. (1998) also gave the example of sand encroachment into 

the cultivated areas of south Kuwait that has forced famers to take 35 percent of their total cultivable land 

out of production. Water and wind erosion are also prevalent in the western area of Libya (Jifara Plain and 

Jebal Naffusah) where agricultural development is dominant, and in El-Witia area in the southwestern Jifara 

Plain as observed by the Remote Sensing Center in Tripoli (Ben-Mahmoud, Mansur and AL-Gomati, 2000). 

For example, the area of fertile cultivated soils in El-Witia classified as class 1 has decreased (1986 to 1996) from 

66 000 to 44 900 ha (a 31 percent reduction), whereas the area of soils low in fertility has increased from 97 

000 to 134 000 ha (a 38 percent increase). In addition, the area covered with sand dunes has increased from 

32 000 to 104 000 ha (a 52 percent increase). 


405 


The Nile Valley system comprising around 2.63 million ha represents the most fertile lands in Egypt - 

and probably in the whole region. However, wind erosion and salinity have degraded the soil resource 

(El-Khole i, 2012). Sand encroachment and mobile dunes are now estimated to cover more than 16 percent of 

the total area of Egypt. These conditions have led to active sand encroachment on the fringes of the cultivated 

areas in most areas of the country. An estimated area of 0.76 million ha has been reported to be affected by 

sand encroachment and active dunes, causing reduction in productivity of as much as 25 percent. The annual 

soil loss was estimated to be about 1.0 million ha. 

In Jordan, the flat topography (

13.3.5 | Soil salinization/sodification 

Previous analyses (Hussein, 2001) revealed that 11.2 percent of the region’s soils are affected by various levels 

of soil salinization. Salt-affected soils vary in extent by country from 10-15 percent in Algeria to over 50 percent 

in Iraq. In the United Arab Emirates 33.6 percent of the area is salinized (EAD, 2009). About 50 percent of the 

reclaimed lands in the Euphrates plain in Iraq and Syria are seriously affected by salinization and waterlogging, 

and about 54 percent of the cultivated area in Saudi Arabia suffers from moderate salinization (CAMRE/UNEP/ 

ACSAD, 1996). In Egypt, 93 percent of the cultivated lands are affected by salinization and waterlogging. The 

salt-affected area in Iran has increased from 15.5 Mha in 1960 to 18 Mha in 1980, to 23 Mha in 1990, and to more 

than 25 Mha today (Qadir, Qureshi and Cheraghi, 2008). In the United Arab Emirates, areas along the coast 

sabkha (salt marshes or lagoonal deposits) are considered highly degraded due to high levels (28.8 dS m -1 ) of 

salinity (Abdelfattah, 2012). In the coastal region of the Abu Dhabi Emirate, salinity is more than 200 dS m -1 

(Abdelfattah and Shahid, 2007). 

Saline and sodic soils are influenced by climate, agricultural practices, irrigation methods and policies related 

to land management (FAO, 1997). Low annual precipitation and high temperatures have also contributed to 

problems of salinity. In many countries of the region, where irrigation completely depends on groundwater, 

excessive irrigation has caused the formation of a shallow water table leading to increased salinization and 

degradation of the soil resource base. Yield reduction due to salinization and/or waterlogging amounts to 25 

percent in Egypt, and has led to a complete loss of productivity and abandoned agricultural lands in several 

countries. From the very scattered information on the extent and characteristics of salt-affected soils, salinity 

and sodicity in the region is rapidly increasing, both in irrigated and non-irrigated areas. Salinity, sodicity or the 

combination of both in some countries of the region are seriously affecting productive areas such as the Nile 

Delta of Egypt, and the Euphrates Valley in Iraq and Syria. The situation is further complicated by association 

with problems of waterlogging and high CaCO 3 

(up to 90 percent in United Arab Emirates, Al Barshamgi, 1997). 


The impact of soil degradation on biodiversity has received little attention in the countries of the region 

and there is little information available. Nevertheless, it is estimated that the region is home to one-tenth of 

the recorded plant species worldwide or about 25 000 species of plants. Of these, 25 percent are endemic to 

the region, 10 percent are of medicinal value, and many are a source of food. This indicates the importance of 

the region as a store of genetic resources (Abahussain et al., 2002). The lack of proper conservation practices, 

overgrazing of herds of ruminants, and deforestation for fuel are causing serious losses of plant cover and 

of valuable genetic resources, including below-ground biodiversity that is rarely quantified in this region. 

Proper and sustainable utilization of plant species which yield valuable products could boost incomes and 

help reduce poverty amongst nomads and local settled populations. However, thousands of plant species and 

varieties have disappeared, and a further 800 plant species are threatened with extinction (Al-Eisawi, 1998) 

and this loss of plant species is likely to result in changes in soil biodiversity. 

Iran is renowned for having one of the richest plant reserves in the world. The country has some 12 000 

species of plants, the majority of which are endemic (the Iranian National Action Programme). In the Elmalha 

area of Sudan, Bakheit (2011) studied the availability and distribution of famine foods and their role in times 

of famine. The study revealed endemic species that are considered as alternative foods in time of crisis but 

which are threatened by genetic erosion due to soil degradation. Soil degradation studies in South Sudan 

indicate the disappearance of palatable grasses such as Panicum turgidum and appearance of less palatable 

grasses such as Aristidia funiculata. The alien species now covers 40 percent of the pasture area, resulting in 

disappearance of many wild animals and decrease in biodiversity (Elfaig, Ibrahim and Jaafar, 2015). The study 

pointed out that drought, unsustainable use of forest and pasture, and increase in population pressure were 

the main causes of this environmental degradation. 


407 


In Tunisia, decades of open grazing in the Bou Hdma National Park have caused severe loss of perennials and 

increased density of annuals (Belgacem, Tarhouni and Louhaichi, 2013). The study reported that grazing has 

reduced total plant cover by 38 percent and the contribution of perennials by 72 percent, while annuals were 

affected 100 percent. Overgrazing of rangelands generally causes replacement of highly palatable species with 

less desirable plants. Along the sea coast of Egypt, overgrazing during the period from 1974 to 1979 decreased 

the total density vegetation by more than 15 percent, mainly due to a decrease in some perennial herbs, while 

at the same time the total cover increased by about 38 percent, due mainly to perennial shrubs and succulents. 

Also in Egypt, vegetation on Mount Elba and the surrounding valleys has been reduced, and the resulting 

increased runoff threatens the diversified natural plant communities in the valleys. Some of these plant 

species are considered to be of high value as genetic resources as they are adapted to the desert conditions. In 

Lebanon biodiversity is threatened by many factors, chief among them are erosion, urban development and 

overgrazing resulting in dominance of xerophytes at the expense of other species (Zahreddine et al., 2007). 

Rising levels of poverty in the Ramallah area of Palestine have led to most farmers (83 percent) turning to the 

collection of medicinal plants for commercial use (Abu Hammad and Tumeizi, 2012). 

13.3.7 Waterlogging 

Waterlogging is a common constraint in irrigated areas of the region because of inadequate drainage. The 

problem is exacerbated by the dominant heavy textured alluvial soils and by seepage from the conveyance 

canals. Soil salinity, sodicity and water logging conditions have definite adverse impacts on soil productivity, 

estimated to be of the order of 30-35 percent of the potential productivity. In many areas of the old Nile Valley 

in Egypt, waterlogging has led to increased soil salinity and in certain areas to increased soil sodicity. In the 

Siwa oasis, for instance, the rate of water table rise during the period 1962–1977 was 1.33 cm yr -1 . Subsequently, 

the rate increased to 4.6 cm yr -1 and consequently subjected fertile soil to degradation (Misak, Abdel Baki and 

El-Hakim, 1997). 

Waterlogging has also become a serious problem on many farms in the United Arab Emirates due to poor 

drainage caused by the presence of a strong and thick hardpan and by excessive use of irrigation water. In 

addition, sea water intrusion in many areas reaches the surface and causes complete vegetation failure 

(Abdelfattah, Shahid and Othman, 2008). In Tunisia, of the 410 000 ha of irrigated area, about 87 000 ha (22 

percent) are affected in varying degrees by waterlogging. This hydromorphy affects most of the irrigated areas 

in the valley of Medjerda, from Ghardimaou up to Kalaat Andalous, and also affects the majority of oases. 

Overall, waterlogging affects 29-67 percent of irrigated areas in the north, 35 percent in the oases of Kibili and 

Toezure, and to a lesser extent the plains of Dorsal and irrigated areas of Gabes and Cap Bon (14-20 percent). 

It also affects some irrigated areas in the far north (Nefza, Sejnane and Mateur) and some irrigated areas of 

the centre. 

13.3.8 | Nutrient balance change 

The problem of nutrient-constrained agriculture is particularly acute in the region. It is associated with land 

use pressure and the consequent intensification of cropping systems and related soil degradation. Nutrient 

depletion is increasingly affecting land productivity in the region. In Sudan, continuous cultivation over nearly 

a century has decreased the base saturation percentage by 25 to 42 percent, indicating leaching with irrigation 

water of soluble anions and cations down the soil profile. Soil degradation due to nutrient depletion in Sudan 

is largely concentrated in the arid and semi-arid parts, particularly in southern Kordofan and Darfur, and in the 

dry sub-humid and moist sub-humid zones of south-western Sudan. This soil degradation is clearly related to 

agricultural activities and to deforestation. 


408 



Soil compaction and crusting are the most serious forms of physical degradation affecting several irrigated 

areas in Libya, especially in sandy soils (Ben-Mahmoud, Mansur and Al-Gomati, 2000). Most soil compaction 

in the region is caused by tillage practices. For example in Iran continuous tilling over more than 50years 

has exposed surface soil to water run off due to an increase of up to 33 percent of soil bulk density. Surface 

compaction and crusting in the Arabian Peninsula, especially in the United Arab Emirates, is often due to land 

filling and levelling for infrastructure development. One study found that compaction increased from a bulk 

density of 1.15 to 1.66 g cm -3 (e.g. by 44 percent) and water infiltration dropped from 267 to 52 mm h -1 (e.g. 

by 81 percent). Soil compaction due to extensive tillage operations in the furrow slice (to 30 cm) is a major 

physical degradation in soils with high clay content where heavy mechanization is practiced. An example is 

documented (Biro et al., 2013) in rainfed farming systems of eastern Sudan, where three decades (1979-2009) 

of cultivation have increased compaction in the 0-5 and 5-15 cm depth from 1.33 and 1.42 g cm -3 in woodland and 

from 1.37 and 1.56 g cm -3 in fallow land to 1.56 and 1.72 g cm -3 (e.g. 16 percent). These levels of compaction have 

contributed to the general decline in productive potential in Sudan. This increase in compaction is comparable 

to the increase in bulk density of 13 percent documented in Jordan during the half century following conversion 

of forest land to cultivation of wheat and barley (Khresat et al., 2008). 

Military activities in the Al Salmi area on the western border of Kuwait have resulted in huge disturbance 

and caused soil and vegetation degradation (Al-Dousari, Misak and Shahid, 2000). The soil pores in the area 

have sealed and the infiltration rate has declined by 19.5 to 64.4 percent. Bulk density has increased by 26-33 

percent. In the Kabd area southwest of Kuwait City, pressure on land has resulted in compaction 20 percent 

higher in non-protected areas (bulk density of 1.8 g cm -3 ) than in protected areas (bulk density of 1.5 g cm -3 ; 

Misak et al., 2002). More generally, it has been estimated that a wide range of activities in Kuwait – grazing, 

quarrying, camping, and agricultural and animal production – have increased compaction by 12.9 to 23.4 

percent (Al-Awadhi, Al-Helal and Al-Enezi, 2005). 

13.3.10 | Sealing/capping 

The population of the region is approximately 6.2 percent of the world population. The region’s fragile 

ecosystem is endangered by one of the highest rates of population increase (3 percent) in the world. This puts 

enormous pressure on the capacity of land resources to provide goods and services. Encroachment of human 

settlements on scarce good quality agricultural land or in areas of adequate rainfall for agriculture occurs in 

many countries of the region, jeopardizing the role of land as a source of food. For example, in Egypt the net 

population density in towns is more than double the recognized maximum threshold of 360 ha -1 . During 1987- 

2007, the cultivated land in the Delta and Nile Valley did increase (to about 7 260 000 ha), but at the same time 

human settlement and land allocated to roads and irrigation canals and drains also increased (by 33.6 percent 

and 40 percent, respectively). As a result, recent studies (ESCWA, 2007) have shown that urban encroachment 

on highly fertile agricultural land in Egypt is emerging as a significant problem. For example, in El-Mahalla El- 

Kobra in the Gharbiya Governorate, the rate of urbanization from 1950 to 1987 was 10 percent annually, but 

from 1987 to 1995 the rate shot up to 33 percent ayear. In the 1950-1987 period, annual loss of agricultural land 

averaged 0.4 percent but it has subsequently risen considerably. 

Iran has the largest urbanized area in absolute terms in the region, followed by Saudi Arabia and Iraq, while 

the highest Urbanization Index is recorded for Gaza Strip, Bahrain, Palestine, Israel and Lebanon (Figure 13.2). 

Land sales also play an important role in the decline in the area of productive lands. In Jordan, the 

agricultural sector lost about 24.3 percent of its land during the period from 1997 to 2007. Rainfed cultivation, 

which represented 89 percent of total cultivated land in 1983, had lost 22.6 percent of its area by 1997. 


409 


The main causes of soil problems in Jordan are: (i) improper farming practices, such as failure to use contour 

ploughing, or over-cultivation of the land; (ii) overgrazing; (iii) the conversion of rangelands to croplands in 

marginal areas where rainfall is insufficient to support crops in the long term; and (iv) uncontrolled expansion 

of urban and rural settlement at the cost of cultivable land. 

Urban populations are growing at 8 percent a year as opposed to just 1 percent in rural areas. In some 

countries of the region nearly the whole population is urban (e.g. Kuwait, 97 percent; Bahrain, 90 percent; 

Saudi Arabia, 83 percent; and United Arab Emirates, 84 percent). This high rate of urbanization has been 

accompanied by conversion of agricultural lands into urban areas. In Libya, over 25 percent of highly fertile 

lands have been taken over by the expansion of urban areas. 

The rapid urban population growth in the region increases the pressure on the natural resources (AOAD, 

2004). An example of dramatic urban expansion is found in Lebanon where a study by Darwish and Khawlie 

(2004) showed that during the period from 1962 to 2000, urban areas expanded by 208 percent while 

agricultural lands decreased by 35 percent. Much of the area converted to settlements was highly productive 

agricultural land on Fluvisols, Luvisols and Cambisols. Some 32 percent of class 1 (prime land) and 26 percent of 

class 2 land were converted into urban areas. 


0 2000 4000 6000 8000 10000 

Saudi Arabia 

Kazakhstan 

Iraq 

Algeria 

Egypt 

Syria 

Morocco 

Libya 

Tunisia 

Israel 

Jordan 

Yemen 

Lebanon 

United Arab Emirates 

Kuwait 

Qatar 

West Bank 

Oman 

Bahrain 

Gaza Strip 

Western Sahara 

20 

15 

10 5 0 


Figure 13.2 Extent of the urban areas and Urbanization Indexes for the Near East and North African countries. Source: Schneider, 

Friedl and Potere, 2009. 


410 


13.4 | Major soil threats in the region 

Most of the land area of the region falls in the hyper-arid, arid and semi-arid climatic zones. About 87 percent 

of the region is predominantly desert. Major soil threats are: erosion, salinity/sodicity, pollution, and soil C 

loss. Main causes of soil degradation are: mismanagement coupled with poor policies; use of inappropriate 

technology; increased levels of traffic movements and road construction; industrial activities and mining; 

urban expansion; deforestation, overgrazing and inappropriate cultivation practices; and dumping of 

hazardous wastes. 

13.4.1 | Water and wind erosion 

Wind and water erosion in the region are due to a complex interaction of factors related to the resilience 

of land resources, land use and management, and socio-economic conditions. In this section, a link is made 

to the prevailing land uses in the different countries of the region, and the consequences and responses are 

discussed. 

Erosion caused by rainfed farming 

FAO (2004) reported that the annual soil loss in Iran due to erosion is 1-2 billion tonnes yr -1 , and that 76 

percent of the total area is under erosion threat. The area affected by wind erosion in Iran is about 12 percent 

of the total country surface area, six times the global rate of 1.8 percent. Other studies also report soil erosion 

as a serious problem in Iran, up to 20–30-times acceptable levels (Jalalian, Ghahsareh and Karimzadeh, 1996). 

In the north-west of Yazd in the Yazd-Ardakan plain, total soil mass transported was measured at 220.93 kg 

m -1 yr -1 and soil loss at 1.356 kg m -2 or about 13.56 tonnes ha -1 yr -1 causing topsoil C reduction at an average 

rate of 4 percent month -1 (Azimzadeh et al., 2008). The Iranian National Action Programme (NAP) reported in 

2004 a lack of access for farmers to inputs as an additional cause of soil erosion. This is due to various driving 

forces like poverty, lack of security and awareness, inadequate extension, absence of technical knowledge and 

financing. 

Elsewhere, driving forces for erosion processes were reported to be interrelated and to result in different 

degrees of degradation. In the El Bayadh region of Algeria, soils on limestone covers were classified as 

moderately vulnerable, vulnerable and highly vulnerable to degradation as a function of their vegetation cover 

(Belaroui, Djediai and Megdad, 2014). The study describes how the Algerian steppe has in recentyears become 

the scene of an ecological and climatic imbalance. 

More generally in the southern part of the Mediterranean region, overgrazing and cultivation of vulnerable 

land in arid and desert regions have induced severe wind erosion. In Morocco, erosion is a serious agroenvironmental 

threat and was found to cause soil losses generally between 12 tonnes ha -1 yr -1 and 14 tonnes ha−1 

yr−1. In some areas of the Rif Mountains these values reach 30 to 70 tonnes -1 ha -1 yr -1 (Benmansour et al., 2013). 

Using radioisotopes (137Cs), these authors found that the tillage process on sloping lands over the last halfcentury 

had resulted in significant translocation of soils within the field. A study by Dahan et al. (2012) found 

that soil erosion in Morocco is a result of several factors. The most important factors have been: increased 

population pressure on limited natural resources; over exploitation of forestry assets; removal of natural 

vegetation from sloping lands; overgrazing; cultivation of vulnerable lands in arid and desert regions; and 

inappropriate land management, mainly tillage practices. They also reported that water erosion accelerated 

by human intervention is the main cause in Morocco of soil degradation and of the deterioration of water 

quality that it entails. Soil erosion in Morocco affects up to 40 percent of its territory with the total annual soil 

loss evaluated at 100 million tonnes, equivalent to 50 million m 3 annual reduction in dam storage capacity. 


411 


In El Bayadh region of Algeria, the Sirocco (a hot, dry wind blowing northwards from the Sahara) with a 

speed of 1.1 to 2.9 m s -1 is a dust-blowing wind which causes significant wind erosion (Belaroui, Djediai and 

Megdad, 2014). A recent study (Houyou et al., 2014) showed that 20 million ha of steppe lands faced a high 

risk of wind erosion due to low rainfall and poorly rooted vegetation on sandy soils. This area is threatened by 

yet more intensive erosion because of newly adopted government policies favouring extensive rainfed cereal 

cropping. The study showed that this decision has led to very high erosion rates (74.4 Mg ha -1 yr -1 ). Clearly this 

cropping system is unsustainable and the policy should be revised. 

In Tunisia, overall soil loss due to water erosion has been estimated to be equivalent to 23 000 ha yr -1 in 

the isohyets above 200 mm. In some areas of Syria, soil loss due to water erosion has been estimated to range 

from 10 to 60 kg ha -1 (under forest), from 200 to 2 550 kg ha -1 (under burned forest) and as high as 960 to 3 280 

kg ha -1 (under agricultural land). High population growth 1984-1999 (4.0 percent annually against a national 

average of 3.0 percent) in the village of Im Mial in the north-west of Syria led farmers to practice continuous 

rainfed barley production with no fallow or rotation (Nielsen and Zöbisch, 2001). This resulted in a very sparse 

vegetation and extensive wind erosion. 

Grazing and tillage practices generally contribute to vulnerability to water erosion. In one study in NENA, 

the highest soil loss was recorded in over-grazed mountains, which lost up to five times more soil than slopes 

under managed grazing (Shinjo et al., 2000). However, soil loss on tilled slopes can be one to four times that 

from grazed areas. Estimates of areas under serious water erosion in northern Iraq showed an increase from 12 

percent in 1954 (Gibbs, 1954) to around 22 percent in 1997 (Hussein et al., 1998), suggesting a rate of increase of 

0.2 percent yr -1 . The main reason was mismanagement of cropland and rangeland during the intervening four 

decades. In Yemen the surface runoff to the sea measured in some major wadis is estimated at 1430 million m -3 

yr -1 (Al-Hemiary, 1999). 

In Jordan, water and wind erosion are both problems. Water erosion in the more vulnerable areas of the 

country with annual rainfall more than 400 mm and a slope greater than 25 percent – occurs on just 2.5 percent 

of the total area of the country. However, water erosion also occurs in the Badia region with its very low rainfall. 

The Badia soils are subject to water erosion because they are bare and highly exposed to what rainfall there 

is. The consequent formation of a slowly permeable seal and crust has enhanced runoff and water erosion 

(Rawajfih, Khersat and Buck, 2005). In Palestine it has been demonstrated that soil conservation pays – net 

profit was found to 3.5 to 6 times higher than without conservation measures. However, farmers’ willingness 

to adopt conservation measures was influenced by other factors too, including knowledge and perception, 

land tenure, and the type of landscape (Abu Hammad and Bǿrresen, 2006). 

Early studies estimated the risk of water erosion in Lebanon to be 50–70 tonnes ha -1 yr -1 , but subsequently 

this has increased to 150 tonnes ha -1 yr -1 (Bou Kheir, Cerdan and Abdallah, 2006) and may be as high as 317 

tonnes ha -1 yr -1 . Some farmers in Lebanon are nonetheless reluctant to change their cultivation practices, 

particularly ploughing up and down slopes of more than 20 percent. Farmers are also expanding cultivation 

on steep slopes, even though they may be aware that this aggravates water erosion. They justify up and 

down ploughing as necessary for tractor performance, and steep lands are the only available land for 

extending cultivation (Zurayk et al., 2001). In general, human activities have been identified as causes of water 

erosion in Lebanon including: encroachment into agricultural land, cultivation of fragile soils, over-grazing, 

deforestation and overexploitation of woodland resources, uncontrolled use of fire for agricultural and forest 

clearing, unsustainable agricultural practices, poor irrigation practices and inefficient water use, and chaotic 

urban sprawl into fertile lands and forests (NAP, 2003). 


412 


In Sudan, studies showed that in areas affected by water erosion, about 74 percent of respondent households 

are exposed to food shortages that are sometimes severe (Akuot Gareng Apiu Anyar, 2006). The expansion of 

mechanized rainfed agriculture (1970–1990) at the expense of rangelands and forests led to land degradation 

by enhancing soil erosion. Subsequently crop yields declined sharply due to the decreased fertility. A principal 

driver has been the rapid growth of the population (2.6 percent per annum -1 ) which increased demand for 

agricultural land. Mechanized rainfed agriculture expanded from 500 ha in the 1940s to 2.3 million ha in 2003. 

In the coastal part of Sudan, although total annual rainfall is low (75 mm), the sandy texture of the soils makes 

them very highly erodible (Elagib, 2011). With the observed increasing seasonality and intensity of rainfall, high 

runoff and erodibility in these areas could be expected to cause heavy soil degradation through soil loss. 

Studies on factors contributing to wind erosion in Sudan showed that in rainfed agricultural zones, deep 

ploughing and leveling of the surface soil caused an increase in its susceptibility to wind erosion, which, in 

turn, has led to a severe decline in its fertility and, in some places, to the formation of sand dunes. The fragility 

index (degraded land in ha divided by population) is a good measure of the extent of growing pressure in fragile 

ecosystems. In Sudan, this value is very high in the hyper arid zone (31.1 percent), the arid zone (30.5 percent), 

and the semi-arid zone (22.5 percent). The value is low in the dry sub-humid zone (7.9 percent) and in the moist 

sub-humid zone (8 percent) (Ayoub, 1998). 

Erosion caused by other land uses 

In Lebanon, increasing demand for construction materials has led to extensive unregulated mining 

activities, including a large number of open quarries. Darwish et al. (2011) reported that, during the period 

from 1989 and 2005, quarries increased by 63 percent, covering an area of 5 267 ha. Many quarries were 

established on sloping lands (62.2 percent), triggering acceleration of water erosion processes. The high 

population density in countries like Kuwait (120 person km -2 ) has a profound influence on soil disturbance 

through uncontrolled human activities. The Nabkha (stabilized aeolian landform developed as result of the 

deposition of wind-driven sediments around desert shrubs) along the coastal plain in Kuwait is used as a land 

degradation indicator (Khalaf and Al-Awadhi, 2012). The average annual sand drift rate in Kuwait is about 20 

m 3 (m width) −1 yr−1 and negatively affects farms causing adverse environmental and economic impacts 

(Khalafa and Al-Jjimi, 1993). Al-Awadhi and Cermak (1998) calculated the average sand movement in Kuwait as 

7.8×104 kg (m width) -1 yr -1 . In the Jalal-Alzor (Kuwait), human activities such as the unregulated use of off-road 

vehicles has resulted in soil disturbance and accelerated soil erosion. There has been a consequent increase 

in the rate of deflated sand from 220.5 kg m -1 width in 1989 to 400 kg m -1 in 2007, a total rate of increase of 81 

percent in two decades (cited by Al-Awadhi, 2013). 

Grazing was found to enhance soil loss by water. In the Matash Mountains of the Alborz Mountain range 

in Talesh Region, Iran (slope of 16 percent and 1 286 mm of precipitation per annum), soil loss due to open 

grazing was more than 26 times that in rainfed agriculture (Sadeghi et al., 2007). In Sudan the dominant 

type of housing is the traditional hut made from forest products. These buildings need to be renewed every 

twoyears on average. A study found that this practice exacerbates the process of soil erosion. Conserving and 

restoring the vegetation cover in these areas was achieved through adoption of mud huts. As a result, the 

stocking density of trees around the villages went up compared to the control (FAO, 2013b). 

Deforestation is also a factor inducing soil erosion. Fuel wood or charcoal production for domestic use is 

one element in this deforestation. Deforestation for agriculture in semi-arid lands around settlement areas is 

also a cause of soil degradation, for instance in the Jifara Plain of Libya. 


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Dust storms 

Dust storms are frequent in the region and widely reported. A dust storm carries toxic elements like Pb 

with concentrations as low as 20 to 288 mg kg–1 in Oman to higher levels of 742 mg kg–1 in Bahrain and 1762 

mg kg–1 in Saudi Arabia (Madany, Akhter and Al-Jowder, 1994; Al-Rajhi, Seaward and Al-Aamer, 1996). The 

problem of sand drifting and dune migration is of special concern in some countries of the region such as 

Saudi Arabia, where approximately one-third of the country is covered by moving sand. Al-Harthi (2002) found 

that these storms have resulted in dune movement of 9.9 m yr -1 (for 4.9 m-high dune) up to 16.5 m yr -1 (for 1.9-m 

high dunes). This problem is exacerbated by human activities such as overgrazing or other activities that may 

destroy the desert pavement which protects the loose sand underneath. 

In Kuwait, the sand drift potential was found to be as high as 354 vector units (Al-Awadhi, Al-Helal and 

Al-Enezi, 2005). Sandstorms are very frequent in summer especially when the wind speed exceeds 6 m s -1 . An 

annual amount of sand drift can measure 7.8×104 kg m -1 width (Al-Awadhi and Misak, 2000). 

As cited by Goudie and Middleton (2001), estimates of rates of dust deposition exist for a number of sites at 

varying distances from the Sahara. The dust originates from southern Algeria, the Nubian Desert in southern 

Egypt and Northern Sudan. Volumes carried to Western Europe are less than 1 g m -2 . Up to 5.1 g m -2 may reach 

Spain, while over Sardinia, Corsica, Crete and the south-east Mediterranean, most values are between 10 

and 40 g m -2 . Long-range transport of Saharan dust to the central Mediterranean is characterized by events 

lasting two to four days, compared to an average duration of just one day for events reaching the Eastern 

Mediterranean from the Arabian Desert (Dayan et al., 1991). 

Several studies reported the frequency of dust storms in the Arabian Peninsula and found that the average 

quantity of dust falling on Kuwait and Riyadh were 191 and 392 tonnes km–2year–1 (cited by Ibrahim and El- 

Gaely, 2012). A recent study (Jish Prakash et al., 2014) on the impact of dust storms on the Arabian Peninsula 

and the Red Sea reported that strong winds (velocities exceeding 15 m s−1) entrained large quantities of 

dust particles into the atmosphere with sources including the lower Tigris and Euphrates in Iraq, areas of 

Kuwait, Iran and the United Arab Emirates, and the basin of the Arabian desert (which includes the Rub’ al 

Khali, An Nafud and Ad Dahna). The study also reported that the frequent dust outbreaks and dust storms 

each year in the NENA region have profound effects on all aspects of human activity and natural processes. 

The total amount of dust generated by the storms is estimated at 93.76Mt, of which 80 percent is deposited 

within the area, around 6 percent (5.3 million tonnes) is deposited in the Arabian Sea, the Gulf received 15 

percent (1.2 million tonnes), and the Red Sea roughly 6 million tonnes. In the Middle East, more than 60 dust 

storms occurred during the period 2003–2011 with significant impact on the countries of the region (Hamidi, 

Kavianpourl and Shao, 2013). Some countries are worst affected. Iraq, for example, experiences on average 

about 122 dust storms and 283 dusty days eachyear. Some experts expect this may increase to 300 dusty days 

and dust storms a year within the next ten years (Kobler, 2013). 

Consequences of soil erosion 

Erosion processes remove the fertile part of the soils and thus reduce the effective depth to be exploited by 

roots and the amount of water available to plants. This is considered a major constraint limiting productivity 

in Morocco (Dahan et al., 2012). In the Maghreb region (Morocco, Algeria, Tunisia), uncontrolled runoff from 

terraces has reached the stage of gully formation. More generally, sheet and gully erosion has become common 

due to increased population, deforestation, overgrazing, and expansion of cultivation on steep land (Dregne, 


414 


2002). There are few studies on the effects of erosion on land productivity in the region. However, one research 

study on barley in Aridisols of Egypt (Afifi et al., 1992; Wassif, Atta and Tadros, 1995) found a declining yield 

of 2 kg ha -1 Mg -1 of soil erosion which is equivalent to 0.21 percent Mg -1 soil erosion. A study in Iran found soil 

erodibility by water was negatively correlated with wheat yields (Vaezi, 2012). The agricultural productivity 

of oases in countries like Sudan and Egypt is threatened by the adverse impacts of sand encroachment and 

mobile dunes. 

Responses to soil erosion 

Ways to contain erosion are very diverse and location-specific. In the sandy depression of El-Farafra, 

Egypt, which suffers from wind erosion, Sallam, Elwan and Rabi, (1995) reported that mixing sandy soils with 

grey shale at a ratio of 15 percent (w/w) improved the quality of these soils. Organic manures, compost and 

synthetic soil conditioners have been used to contain wind erosion. Compost alone or compost combined 

with hydrogel conditioners was found to decrease erosion by 58 to 74 percent (El-Hady and Abo-Sedera, 2006). 

One study found that water erosion could be stemmed in Morocco by increasing levels of C in the topsoil 

using conservation measures (Mrabet et al., 2001). Irrespective of rotation, conservation measures were 

found to increase topsoil C by 44 percent. The study concluded that systems with increased C are generally 

characterized by diminished erosion. 

Government policy responses can play a vital role. For example, public policy in Egypt and Iraq has been 

determinant in increasing green cover in desert soils in those countries (Nielsen and Adriansen, 2005). The Iraqi 

case is a negative one, illustrating the effects of deliberate government policies in draining the marshlands, 

which resulted in significant land degradation. Recent efforts have been devoted to the re-establishment of 

these marshes. They are being re-flooded and vegetation is returning. The Egyptian government promoted 

land reclamation after the 1952 revolution aiming at increasing agricultural production. This reclamation 

was executed through internationally funded developmental projects and with local funding and involved 

distributing small areas of lands to graduates. 

Efforts at containing water erosion in rainfed farming in Yemen depend on the establishment and 

maintenance of terraces as conservation structures, a highly labour intensive task. However, labour shortages 

and lack of profitability have been constraints, and the degradation of these structures has continued. In 

Jordan, the use of polyacrylamide (PAM) at application rates of 10 to 30 kg ha -1 was found to be very effective 

in reducing runoff and soil loss by up to 23 percent and 53.9 percent, respectively. As a result, dry matter crop 

yield went up by 35 to 56 percent (Abu-Zreig, Al-Sharif and Amayreh, 2007). Interestingly, this study developed 

an increased threshold runoff value of 0.56 mm rainfall in control sites to 1.11 mm in PAM plots (e.g. close to 100 

percent). Also Abu-Zreig (2006) pointed out that PAM with more surface area (30 percent) reduced soil loss by 

approximately 46 percent compared to the 24 percent reduction with less surface area (20 percent). 

Soil conservation to mitigate erosion has also been done by planting trees and grass along wadis. 

Construction of diversion banks and dams across watercourses has been carried out on slopes and stream 

beds in Libya. One study (Mohammed, Stigter and Adam, 1996) found that windbreaks reduced wind speed by 

about 20 percent and in turn limited sand deposition. However, sand deposition also occurred within the belt 

which after some time started to act as a zero permeability wind break. This study suggested that control of 

sands in the source area using shelterbelts may not be sufficient in the long term. 

Some countries have introduced policies to reverse land degradation due to erosion. In the mechanized 

rainfed projects of Sudan, farmers are required to leave 10 percent of the cultivated area for forest trees. To 

protect soil against erosion after fire in Lebanon, a ministerial decision (181/98) imposed a five year ban on 

grazing on public land after fires to enhance land cover recovery. 


415 


13.4.2 | Soil salinization/sodification 

Distribution of salt-affected soils in the region varies geographically with climate, agricultural activities, 

irrigation methods and policies related to land management. These soils are mainly confined to irrigated 

farming systems in the arid and semi-arid zones. The salts present are either of intrinsic origin (typical of Egypt, 

Sudan and Iran) or are the result of sea water intrusion in coastal regions or of irrigation with brackish or 

saline groundwater. In the irrigated zones of Morocco, continuous irrigation has resulted in soil salinization. 

Secondary salinization due to irrigation with saline water is also reported in the NENA region. In Libya, Sudan, 

Iran, Iraq and United Arab Emirates, large tracts of lands have been degraded due to heavy irrigation with 

groundwater. Salinity, sodicity or the combination of both are seriously affecting productive areas like the Nile 

Delta of Egypt and the Euphrates Valley in Iraq and Syria. The situation is further complicated by association 

with problems of waterlogging and high CaCO 3 

(up to 90 percent in the United Arab Emirates, Al Barshamgi, 

1997). In Kuwait and the United Arab Emirates, soil salinization is mainly confined to coastal areas, but also 

occurs on irrigated farms. 

Local soil conditions can worsen the situation. For instance in the southern part of the Jordan Valley, the soil 

is characterized by high salt content, poor permeability and high gypsum content. The degradation is worse 

when low quality irrigation water, for example treated waste water, replaces fresh water. In Libya, El-Tantawi 

(2005) reported that the soils of the Jifara Plain are usually calcareous and often shallow, with huge areas of 

calcrete outcrops developed during the Pleistocene epoch. Gypsum encrustation is commonplace in the drier 

parts where annual precipitation is below 200 mm. 

Salinity problems in the region also stem from inadequate irrigation water management (Al-Hiba, 1997). 

In almost all countries of the region with coastlines, heavy extraction of groundwater has led to intrusion of 

seawater into aquifers, thereby raising the content of salts in the water. An example is the Batinah aquifer of 

the Sultanate of Oman where seawater is intruding at an alarming pace (Naifer, Al-Rawahy and Zekri, 2011). 

The cause was the expansion of agriculture since the 1980s which accelerated the overuse of groundwater, 

disturbing the water balance and ultimately leading to water intrusion from the sea. In the Jifara Plain of Libya, 

increased human pressure on aquifers has induced seawater intrusion in the coastal zones and a combination 

of over-irrigation and inefficient drainage causes waterlogging and secondary salinization. 

The main causes of build-up in salinity in the region are: (1) improper functioning or absence of drainage 

systems; (2) a rise in groundwater salinity combined with high rates of evapotranspiration; and (3) high salinity 

in irrigation water. Siadat, Bybordi and Malakouti (1997) recognized natural factors and humaninduced factors 

that cause salinity in Iran. The natural causes of soil salinity in Iran are geological conditions, climatic factors 

(evaporation, rainfall and wind), salt transport by water, and intrusion of saline bodies of water into the 

coastal aquifers. However, of greater concern and importance is humaninduced salinity. This type of salinity 

can stem from a number of causes, including: poor water management, over-grazing, improper land levelling, 

and overuse of groundwater leading to saline water intrusion. 

Secondary salinization due to irrigation with saline water is also reported in the region. In the Tadla irrigated 

perimeter in Morocco, soil degradation through secondary soil salinity and sodicity are caused by heavy 

irrigation with ground- and surface water together with agricultural intensification. In Algeria, secondary 

salinization affects 10 to 15 percent of the total irrigated land. About 90 percent of the agricultural farms in 

Al Ain in the United Arab Emirates are affected by salinity (Abdelfattah, Shahid and Othman, 2009). It is clear 

that in some countries of the region, secondary salinization due to irrigation does not develop. One example 

is in the Vertisols of Sudan, despite more than 80years of irrigation. This is mainly due to the low salt content 

of Nile water (EC of 120 to 220 μS cm -1 ). 


416 


Consequences of salinization 

Salinity in the region has badly affected cropping systems and in many cases has significantly reduced crop 

yields. For example, soil salinity in the Jifara plain in Libya has caused wheat yields to decrease from 5 tonnes 

ha -1 in the 1980s to just 0.5 tonnes ha -1 by 1987. In Iran the annual economic losses due to salinity are estimated 

at more than US$ 1 billion (Qadir, Qureshi and Cheraghi, 2008). The coastal area is one of the most highly 

populated regions of Oman, especially the Batinah area where about 52 percent of land is under cultivation 

and suffers from salinity. Naifer, Al-Rawahy and Zekri, (2011) showed that when salinity increases from low 

(less than 2.5 dS m -1 ) to medium (7.5 dS m -1 ) and to high (more than 7.5 dS m -1 ) levels, losses are equivalent to 

US$ 1 604 ha -1 and US$ 2 748 ha -1 , respectively. Soil salinity, sodicity and waterlogging conditions have definite 

adverse impacts on soil productivity, in the range of 30-35 percent of potential productivity. 

Responses to salinization 

There are many responses in the region to contain the salinity threat such as: (1) direct leaching of salts, (2) 

planting salt tolerant varieties, (3) domestication of native wild halophytes for use in agro-pastoral systems, 

(4) phytoremediation or bioremediation, (5) chemical amelioration, and (6) the use of organic amendments. 

In Iraq and Egypt, surface and subsurface drainage systems have been installed to control rising water tables 

and arrest soil salinity. In Iran, Syria and other Gulf countries, crop-based management and fertilizers are used 

to combat salinization (Qadir, Qureshi and Cheraghi, 2007). In Iran, Haloxylon aphyllum, Haloxylon persicum, 

Petropyrum euphratica and Tamarix aphylla are potential species for saline environments (Djavanshir, 

Dasmalchi and Emararty, 1996). Atriplex is a fodder shrub adapted to arid lands which can bring annual income 

as high as US$200 ha -1 (Koocheki, 2000; Tork Nejad and Koocheki, 2000). Breeding salt tolerant varieties of 

crops (e.g. wheat, barley, alfalfa, sorghum) is also a response to saline environments, although most results so 

far are based on controlled environments rather than on actual yields from the field. 

The use of organic amendments in Egypt showed that the mixed application of farmyard manure and gypsum 

(1:1) significantly reduces soil salinity and sodicity (Abd Elrahman et al., 2012). Recently, phytoremediation or 

plant-based reclamation has been introduced in the region. In Sudan there are very good responses for control 

of sodicity relying on phytoremediation, superior to results from the gypsum amendment traditionally used. 

The production of H+ proton in the rhizosphere during N-fixation from some legumes like hyacinth bean 

(Dolichos lablab L.) removed as much Na+ as did gypsum application which indicates its importance in calcite 

dissolution of calcareous salt affected soils (Mubarak and Nortcliff, 2010). 


Information on soil carbon changes in the region is scarce. Estimates of soil C sequestration are basically 

confined to the work on the drylands ecosystems of West Asia-North Africa (WANA) carried out by Lal (2002). 

These data are nonetheless very useful and can be applied across NENA. Lal’s study indicated that the total loss 

of soil-C from the WANA region could be about 6 to 12 Pg. Despite the low soil C levels in the region (generally 

less than 5 g kg -1 ), with effective control measures of degraded soils, the region could sequester C at the rate 

of 0.1 to 0.2 Mg ha -1 yr -1 (for irrigated crop land) and 0.05 to 0.1 Mg ha -1 yr -1 for both rainfed and rangeland. In 

other words, with desertification control, reclamation of salt-affected soils, and intensification of agriculture 

on undegraded soils, the soils of the region have the potential to sequester 24 to 31 percent (168-380 Tg yr -1 ) of 

the total global drylands soil C (710-1220 Tg yr -1 ). The potential annual sequestration rate could reach values 

between 0.2 and 0.4 Pg C yr -1 , compared to the 1.0 C yr -1 in total global drylands (e.g. 20 to 40 percent). 


417 


SOC change in rainfed farming systems 

The Century model was used by Poussart, Ardö and Olsson, (2004) in a study of the Arenosols of Kordofan 

(Sudan) to estimate soil C levels and changes with reference to values prior to known human interaction. 

Changes were due to pastoral activities combined with cultivation of rainfed crops like Pennisetum 

typhoideum), sesame (Sesamum indicum), sorghum (Sorghum vulgare), and groundnuts (Arachis hypogaea). 

The base scenario modelled indicates that the land management practices continued for more than a century 

have led to a loss of C of about 180 g m -2 (1.8 tonnes C ha -1 ) which is equal to 1.6 gm -2 yr -1 or approximately 

equivalent to half of the historical level of C in the year 1890. Additionally, Ardö and Olsson (2003) modelled C 

changes in north Kordofan in the top 20 cm during the period 1800–2100, with mostly Arenosol and Vertisols 

soil types. They found that C estimates in cropped land have dropped from 16.64 million tonnes in the year 

1800 to 9.16 million tonnes (e.g. a 45 percent reduction), whereas C in other land uses such as shrublands, 

savannah, grassland, or barren or sparsely vegetated land remained almost constant. Another study in north 

Kordofan found that rapid population growth has caused huge soil C loss (73 percent at rate of 16.9 g C m -2 yr -1 ) 

in the top 0-20 cm from 851 g C m -2 in 1963 to 227 g C m -2 in 2000 (Ardö and Olsson, 2004). 

The effects of cultivation of heavy textured soils on C change in some countries of the Mashreq region 

(Jordan, Syria, Lebanon, Iraq, Palestine) have been found to be broadly similar to results from the Maghreb 

(Morocco, Tunisia, Algeria, Libya). Masri and Ryan (2006), in a study on Syria, compared the effects of more 

than twentyyears of wheat cultivation in rotation with lentil (Lens culinaris), chickpea (Cicer arietinum), 

medic (Medicago sativa), vetch (Vicia sativa) and watermelon (Citrullus vulgaris) with continuous wheat 

grown in montmorillonitic thermic Chromic Calcixert. The trial showed that, apart from the rotation of wheat 

with medic which has higher C (8.0 g kg -1 ), C content in continuous wheat cultivation (6.3 g kg -1 ) or in other 

rotations (6.6-7.0 g kg -1 ) differs little from fallow (6.6 g kg -1 ). This suggests that conventional farming systems 

in the Vertisols of the region may not be depleting soil C. 

In the western part of Jordan, the change in C is a function not only of a land use system but also of the 

human impact in a specific land use. For example, cultivating tobacco and clearing residues, and growing 

irrigated cereals retain almost similar amounts of C (6-7 g kg -1 ), whereas ploughing rainfed cereals and 

grazing with compaction consequences retain C contents of only about 1.1 g kg -1 (Khresat et al., 1998). In the 

Lorestan Province of Iran, management practices on rangeland in areas with slopes of 10 to 26 percent caused 

pronounced reduction in C, sometimes twice levels found in dryland farming. When wheat was subsequently 

grown, wheat dry matter, grain yield and grain weight all reduced relative to the control dry farming (by 14, 33 

and 21 percent respectively), which indicates the degrading effects of using these slopes as rangeland (Asadi 

et al., 2012). 

SOC changes in irrigated farming systems 

Irrigated soils in the southern Mediterranean region can be considered as incubators providing optimal 

conditions (humidity and temperature) for microbial activity and thus a rapid degradation of organic carbon. A 

mean annual variation rate of organic matter of -0.09 percent yr -1 over a decade was established in the Doukkala 

region of Morocco (Badraoui, 1998). This decrease of organic matter is attributed to the non-incorporation 

of crop residues into the soils. Crop residues contribute about 30 percent of the total forage consumption 

in Morocco. Countries like Morocco experience rapid organic matter decomposition and turnover and hence 

very low levels of organic matter are converted into humus. The reasons for this rapid decomposition include: 

the high temperature, widespread of use of tillage, clean fallow, overgrazing, no practice of residue recycling, 

and climatic conditions. 


418 


Rates of C change are influenced not only by land use but also by soil type. Clay soils tend to counteract 

decomposition and hence reduce chances of C loss. For example, in clay soils of Morocco, Bessam and 

Marbet (2003) reported that in seven years of continuous tilling, soil SOC reduced in the 0-20 cm profile by 

17 percent (e.g. 0.33 g kg -1 yr -1 ). As cultivation of such soils advances, it was possible over time (an eleven year 

period was monitored) to store more C (by 3.7 percent) in the entire 0-20 cm depth but not the top 2.5 cm 

due to incorporation. The decrease in topsoil C noted increases vulnerability to degradation by erosion due to 

reduction in topsoil buffering capacity or resilience. 

It is apparent that continuous cultivation does not always consume soil C. There are exceptions, however, 

for example the case of the irrigated Vertisols of the Gezira Scheme of Sudan which have been continuously 

cropped for more than 80 years. The carbon content in a furrow slice (0-10 and 10-35cm) of permanent fallow 

(4 and 4.61 g kg -1 ) was lower than that in cultivated plots (6.35 and 5.44 g kg -1 ) by about 37 and 15 percent 

respectively which indicates that C loss due to cultivation in such soils is not a small problem (Elias and Alaily, 

2002). Physical protection of C in heavy textured soils under long-term cultivation (about 30 years) and with 

intensive tillage decreased C in the top 0-15 cm by 15 to 24 percent relative to wood and fallow lands (Biro et 

al., 2013). 

In desert soils carbon could likely be increased by irrigation of alfalafa rotated with wheat or wheat rotated 

with fallow. One study showed this to have increased C in the top 0-10, 10-20 and 20-30 cm depths by factors 

of 5.1-6.8, 3.0-4.6 and 6.8-11.6 respectively (Fallahzadeh and Hajabbasi, 2012a). This rotation could thus stabilize 

soil aggregates and consequently reduce the vulnerability of desert soils to erosion by either wind or water. 

However, C could also be decreased by 24 to 47 percent during bio-remediation (land farming processes) of 

soils contaminated with hydrocarbon of petroleum origin where microbial activities are greatly enhanced due 

to aeration (Besalatpour et al., 2011). 

SOC change and forest clearing 

Tree plantations have been found to be the best system to conserve C in the region. Other land use systems 

practiced on light soils in the region seem to result in C loss. For example, in the sandy soils of western Sudan, 

taking tree C in the 0-30 cm depth as a reference, three years of sole cropping or cropping mixed with trees 

caused about 41 to 47 percent C reduction (El Tahir et al., 2008). 

In Jordan, land resources have been affected by rapid land use change, accelerated by socio-economic 

factors including high population growth, urbanization and agricultural intensification. Farmers converted a 

forest located in Ajloun to cultivation of wheat (Triticum spp.) and barley (Hordeum spp). After more than 50 

years of cultivation, Khresat et al. (2008) reported that soil C in the top 0-20 cm of mostly Inceptisols, Mollisols 

and Vertisols had decreased from 6.73 g kg -1 to 4.70 g kg -1 (e.g. a 30 percent reduction). In Iran, the conversion of 

forest or pasture to arable lands in five cultivation sites where tillage has now been practiced for 40-50 years 

was found to have caused 30-68 percent loss of topsoil (0-20 cm) SOC, specifically from 3.86-9.84 to 3.25-8.06 

kg m -2 (Golchin and Asgari, 2008). Farming practices had also mixed SOC down the profile, leaving less SOC 

on the topsoil for surface protection against degradation by erosion and also reducing the capacity of the soil 

to retain nutrients. 

Also in Iran, change from pasture to dryland farming on Inceptisols was found to have degraded soil and 

reduced SOC by about 67 percent. Recovery times can be very long indeed. For example, a study found that 30 

to 60 years will be required to restore soil C to the initial content of rangeland ecosystems of Chaharmahal and 

Bakhtiari Province in Central Iran. After more than a century of cultivation, SOC in the top 0-15 and 15-30 cm 

had reduced by 25-34 percent (Chigani, Khajeddin and Karimzadeh, 2012). 


419 


The change in SOC depends on soil type. For example, continuous cultivation of Cambisols in northern Iran 

tends to decrease C by 29 to 74 percent as compared to grass or forest use respectively, whereas the equivalent 

decline for Vertisols is only 11 to 59 percent. 

Another study on soil quality indicators potentially sensitive to land degradation due to land use change 

was carried out in a rangeland pasture in Iran’s Isfahan Province that had been protected for more than two 

decades (Nael, Khademi and Hajabbasi, 2004). The study showed that C content in the protected range was 

close to double (1.7 times) that of areas of uncontrolled grazing. This study clearly indicated that in the dryland 

of this region, decisions that control grazing are favourable for soil C storage. 

In another study, four decades after conversion from oak forest (Quercus brantii Lindl.) to vegetable 

cultivation (tomato and snap bean), soils had lost almost 53 percent of C (Fallahzadeh and Hajabbasi, 2012b). A 

further study found that twentyyears after oak forest (Quercus brantii) in the Lordegan region of Iran’s Zagros 

mountains was converted into either wheat or barley cultivation or agroforestry uses, soil C in the 0-30 cm 

depth decreased by 52 and 61 percent, respectively (Hajabbasi et al., 1997). 


Land degradation due to accumulation of contaminants is concentrated in either oil producing countries 

or those that are heavily populated. In agricultural soils, contamination is generally restricted to irrigated 

farming systems. In some instances, the over-use of chemicals (fertilizers, pesticides and herbicides) has 

sharply increased the amount of chemical nutrients in the drainage water, causing water eutrophication 

(NAP, 2002). Since the construction of the Aswan High Dam of Egypt in 1970, fertilizers and pesticides have 

been heavily used in order to substitute for the loss of fertile sediments. FAO (2012) reported that during the 

period from 1950 to 1990, chemical fertilizer use increased more than fourfold, from 2 143 tonnes in the 1950s 

up to 11 700 tonnes in 1990. These chemical fertilizers and the residues of applied pesticides have caused the 

contamination of soil and water resources in the Nile Delta. 

In Iran, contamination of soil is increasing from a variety of sources: petroleum hydrocarbons spilled during 

transportation, leakage from tanks, accidental spillage, pipeline ruptures, or dumping of oil landfill. This 

contamination threatens soil functions. It may decrease seed germination of grasses by more than 50 percent 

(Besalatpour et al., 2008) – although this may not necessarily affect their subsequent performance. It also 

reduces dry matter accumulation in sunflower and safflower by 50 and 73 percent, respectively (Besalatpour 

et al., 2008). The two Gulf wars in 1990 and 1991 contributed to contamination of this kind through the 

detonation of oil wells (Al-Senafy et al., 1997; Misak, Khalaf and Omar, 2009). 

Case study: Kuwait experience in remediation of oil contamination 

After the Iraq war, over 300 oil lakes covering an area of 46 km 2 were formed within Kuwait. The lakes were 

up to two meters deep, and the oil penetrated the soil to varying depths. To restore areas degraded by oil, 

the Kuwait Institute for Scientific Research and the Japan Petroleum Energy Centre began in 1994 to devise 

biological technologies for remediation and rehabilitation (Figure 13.3). 

In a small scale pilot (1 920 m 2 

) and in field demonstrations, heavily oil contaminated soil was remediated 

over a 12-18 month period, using bioremediation techniques involving enhanced land farming techniques, 

windrow composting piles, and static bioventing piles. The programme resulted in 80-90.5 percent 

reduction in the total petroleum hydrocarbons and total alkanes (Al Awadhi, 1996; Balba et al., 1998). This 

technology is considered to be economical, energy efficient, and environmentally friendly with minimal 

residue disposal problems. However, the volatilization of airborne volatile organic compounds in the 

atmosphere during the process of degradation may lead to serious human health risk (Hejazi, Hussain 

and Khan, 2003). 


420 


Figure 13.3 Layout of the project site source (a) and conceptual design and layout of bioremediation system (b). Source: Balba et al., 

1998. 


421 


Contamination of soil by heavy metals is also an issue (Misak, Khalaf and Omar, 2009). In central Iran, 

farmers are extensively using sewage sludge as a fertilizer for vegetable production and, in the absence of 

regulation, heavy metals tend to accumulate in the soil (Afyuni, Rezaeinejad and Schulin, 2006). The so-called 

‘global dust belt’ that extends from the west coast of North Africa, through the Middle East into Central Asia 

(see Section 7.4) transports mineral dust in the region. This dust may carry contaminants and this has been in 

fact the main source of soil pollution with heavy metals in the Arabian Peninsula. The dust was found to carry 

high levels of lead (65 mg kg -1 in Muscat, 742 mg kg -1 in Bahrain and 1762 mg kg -1 in Riyadh, Saudi Arabia) and 

nickel (43 to 3033 mg kg–1 in Muscat) (Yaghi and Abdul-Wahab, 2004). 

In general, soil contamination depends on the distribution of contaminants influenced by high intensity 

rainfall of short duration that results in short runoff, by dust storms, and by human induced factors such as 

mixing residual oil with soil, transport to new areas and dumping in selected sites. 

Responses to soil contamination 

Regional policies for combating desertification 

Most NENA countries have policies and programmes for protection of natural resources, including 

combating desertification. However, many of these policies and programmes (e.g. in Jordan) emphasize 

protective measures, and do not adequately consider rehabilitation or the dimension of the economic and 

social cost of land degradation (Al Karadsheh, Akroush and Mazahreh, 2012). 

Iran has implemented nine strategies for sustainable development. Egypt has strategies for each agroecological 

zone. Lebanon has initiated a large-scale reforestation program and is very active in fighting the root 

causes behind land degradation, mainly by promoting the development of rural areas and reducing regional 

disparities. The national efforts to combat desertification in Oman have concentrated on development and 

conservation of water resources, improvement of land capability and rehabilitation of rangeland. Saudi Arabia 

has programmed an array of activities including capacity building, controlling urbanization, sustainable 

agricultural development, improvement of water sector, legislation, rehabilitation of degraded rangelands, 

forest development and sand dune stabilization. Sudan is integrating strategies for poverty alleviation with 

programmes to combat desertification and these include activities for improvement of land resources, 

production systems and protection of the environment. Syria has implemented many projects aimed at 

expansion of plant cover, controlling desert invasion, establishment of protected areas and green oases, 

sand dune fixation and afforestation. United Arab Emirates has ambitious programmes to improve degraded 

ecological systems, conserve biodiversity, mitigate climate change effects, and combat desertification. 


422 



13.5.1 | Case study: Iran 

Soil nutrient changes 

Change in land use and the use of fertilizers are the main factors affecting soil nutrient change in Iran 

(Shiranpour, Bahrami and Shabanpour, 2011). A study in Gilan Province compared the status of forest soils and 

the same soils turned into tea gardens over a period of 10 to 40years and showed a significant decline in the 

amounts of nitrogen, potassium, phosphorus, calcium and exchangeable magnesium. Deforestation effects 

on soil nutrient losses have been studied in Kajoor watershed in Sari city where results indicated significant 

losses of organic matter and phosphorus. Land uses and different management types were compared in the 

Taleghan area. The study showed the negative effect of irrigation and monoculture on soil nutrients, while 

horticulture and pasture land uses scored better (Sohrabi and Zehtabian, 2012). 

Soil pollution changes 

In Iran, surveys carried out in areas where soil pollution occurs indicate that heavy elements make up the 

majority of soil contaminants. These pollutants originate from a range of sources, including geological and 

mining sources, industrial pollution, petroleum spills, sewage sludge application and excessive usage of 

fertilizers on agricultural soils. 

Losses and sequestration of soil carbon 

Loss of organic matter in soils of Iran is among the most important consequences of soil erosion. Greening 

barren land and improving soil management can significantly increase soil carbon sequestration. Forest 

ecosystems in equilibrium, with both trees and other vegetation cover, are the principal reservoir of organic 

carbon. Varamesh et al. (2010) assessed the effects of reforestation in Tehran Cheetgar Park on carbon 

sequestration and soil characteristics. The study indicated that soil carbon sequestration of Acacia senega.is 

equal to 78 tonnes ha -1 , and of Conifer Species 57 tonnes ha -1 . In general, the carbon sequestration process led 

to improvement of soil and water quality, increased fertility and an improved soil hydrology system as well as 

preventing erosion and reducing nutrient loss. 

Salinity changes 

Yazdani-Nejad and Torabi-Golsefidi (2013) examined the spatial variations and salinity zoning of agricultural 

soil in an area of southern Tehran. About 30.4 percent of the area, covering 20 000 ha, was found to be 

without any salinity problem. These lands were located mainly in areas where irrigation water from deep wells 

was used. A further 42.4 percent of the land had an electrical conductivity of 2 to 4 dS m -1 . In these sections, 

irrigation was with water from deep wells but wastewater was also used for irrigation due to water use 

restrictions. Zones with conductivity of 4 to 8 dS m -1 occupied 22.6 percent of the area, located in the flat plains 

in the southern part. These zones were frequently irrigated with waste water and water from shallow wells, 

and also with low quality water from the downstream sections of the Kan River. Finally, 4.5 percent of the 

land, located in low lying areas, had high electrical conductivity of 8 to 13 dS m -1 . The overall finding of the study 

was that it was the position of the land in the landscape, the depth to the water table and the quality of the 

irrigation water that determined the degree of salinization of the land. Other factors that may play a role are 

the length of the interval between irrigations and the texture of the soils. 


423 


Figure 13.4 Rate of water erosion in Iran. Source: Soil Conservation and Watershed Management Research Institute. 

Water erosion 

In Iran about 40 percent of the country experiences a low erosion rate (less than 10 tonnes ha -1 ), 25 percent 

of the area has a moderate rate of water erosion (10 to 20 tonnes ha -1 ), 23 percent area of the country has a 

high erosion rate (20 to 50 tonnes ha -1 ), and 12 percent has a very high erosion (more than 50 tonnes ha -1 ). 

Figure 13.4 shows the erosion rates in different regions of Iran. 

Results of recent research conducted in the watersheds of the country in recentyears suggest an increased 

rate of erosion according to various models used (Hosseinkhani, 2013; Kavian et al., 2014; Karam, Safarian 

and Hajjeh Froshnia, 2010; Zomorodian and Rahimi, 2012; Bayat et al., 2011; Naderi, Karimi and Naseri, 2010; 

Zare Bidaki and Badri, 2014). Land-use change is one of the most important factors that exacerbate erosion in 

basins (Mohamadzade, Charm and Eskandari, 2014; Ajami, Khormali and Ayoubi, 2012). 

Wind erosion 

In Iran’s deserts and arid areas, rapid changes in temperature cause pressure gradients in different parts 

which result in constant strong winds. Due to these strong winds and to the lack of moisture and vegetation, 

both small and large soil particles are transported, leading to soil erosion and deposition (Mehrshahy and 

Nakoonam, 2009). Sand dunes are estimated to cover 12 million ha. Half of these dunes are active or semiactive 

(Refahi, 2004). The density of air deposits measured over a 40year period shows a growing trend in 

wind erosion in recent years. Sediment textures have also changed, with a significant increase in evaporated 

deposits such as salts and gypsum. 


424 


Studies show that both climatic and human factors play an important role in the development of wind 

erosion. The most important climatic factors are rainfall patterns, rising temperatures, intense evaporation 

from the playa, and reduced intake of water. High wind strengths also reduce the moisture level of the soil 

and enhance wind erosion rates. The human factor is related to land use and concerns the reduction of water 

entering the playa because of dam construction and excessive pressure on pastures and agricultural land in 

recent years (Hosseinzade, Khaneabad and Bargi, 2011). 

Figure 13.5 shows days with dust storms in 2012, while Figure 13.6 shows the origin of dust storms in 2012. 


425 


Turkey 

Iran 

Lebanon 

Syria 

Iraq 

Palestine 

Jordan 

Kuwait 

Dust source areas 

Saudi Arabia 

Bahrain 

Qatar 

Figure 13.6 Internal and external dust sources in recent years in Iran. Source: University of Tehran, 2013. 

The origin of dust in dust storms in Iran is both internal and external to the country. Since 2006 considerable 

dust has affected the west and south west of the country, originating from Iraq and Syria. Other areas such 

as parts of Jordan, Kuwait and northern Saudi Arabia are also involved in the creation of dust in Iran (Jalali, 

Bahrami and Darvishi Bolurani, 2012). Although many recent dust storms in Iran have foreign origins, this does 

not mean that domestic sources play no part in this phenomenon (Fattahia, Noohia and Shiravand, 2012). 

13.6.2 | Case Study: Tunisia 

Between 2006 and 2010 a detailed study was undertaken to assess land degradation in Tunisia. A 

standardized methodology was adopted to map in an interdisciplinary way all aspects of land degradation 

(type, extent, degree, trends and impact on ecosystem services) at a national scale. The exercise involved a 

large number of national institutions and stakeholders, together with international expertise (FAO, 2011a). In 

addition, an inventory was made of successful local practices that combatted land degradation (FAO, 2011b). 

This study was complemented by three investigations at local level (in Kasserine, Siliana and Médenine) that 

refined the identification of the socio-economic pressures and drivers behind land degradation (FAO, 2011c). 

The results have subsequently been expanded and refined. 

Major outcomes of the investigation, relevant for the present assessment of soil change and its impact on 

ecosystem goods and services were: 

1 - The preparation of a national land use system map 

The scale of this map is 1:500 000. It is based on a rasterized database at 30 arc seconds. It was prepared 

using a standard methodology (Nachtergaele and Petri, 2011). The draft map was validated by regional 

institutions in the country and was later refined by simplifying the pastoral classes and by introducing a 

specific unit that concerned alfalfa areas which make up 170 000 ha of the country. 


426 


2 - Land degradation assessment mapping 

The assessment used a standard methodology (Liniger et al., 2011) that allowed the participatory mapping 

of the major types of ecosystem degradation (soil, water and biological) and sub-types (for instance, water 

erosion, compaction, decline in ground water quality or reduction of vegetative cover). At the same time the 

intensity (degree) and the trend of the ongoing ecosystem change was evaluated on the basis of arbitrary 

classes (typically ranging from none to severe or slow to fast). The evaluation was conducted in a participatory 

way involving various stakeholders in the assessment and using hard data where they were available. The 

direct pressures (for instance improper soil management, deforestation or natural causes) and the indirect 

socio economic causes (for instance lack of knowledge or investment) were also determined in the same 

participatory way, as was the impact on ecosystem goods and services. Examples illustrating the outcome of 

the assessment are given in the following maps which evaluate water erosion (Figure 13.7a) and wind erosion 

(Figure 13.7b). Differences in the extent of ongoing degradation processes in the country could also be mapped. 

Figure 13.7 Assessment of Water (a) and Wind Erosion (b) in Tunisia 


427 


Figure 13.8 Soil Conservation in Tunisia 


428 


Figure 13.9 Type of ecosystem service most affected. 


429 


3 - Sustainable land management mapping. 

In each unit of the national map, the prevailing land management practices were inventoried using a 

standard methodology (Liniger et al., 2011). This allowed the characterization of practices in terms of type of 

intervention (agronomic, vegetative etc) and in terms of objectives (prevention, mitigation or rehabilitation). 

At the same time, the extent, trend and efficiency of the conservation practices were assessed. Examples of 

outputs are given for the main conservation types used in the country (Figure 13.8.) and for the impact of soil 

degradation on ecosystem goods and services (Figure 13.9.). 

13.6 | Conclusions 

Although there is a wealth of local and national studies on soil change in the region, a systematic and 

standardized approach is lacking. Results on the extent and intensity of soil change processes still refer to the 

GLASOD study carried out in the late 1980s. 

The degradation of natural resources in arable lands is considered as one of the main threats to agricultural 

production in all countries of the region. Ecosystem service quality and capacity is greatly reduced by degradation 

caused by salinity, erosion, contamination and poor management that leads to a loss of soil organic matter. 

Water erosion is predominant in that part of the region which has sloping lands. Where rainfed agriculture 

is practiced, water erosion may even occur in gently sloping areas. Wind erosion is also a causative factor of 

topsoil removal. Population increase has resulted in soil disturbance due to uncontrolled human activities 

such as mining and open quarries that have triggered and accelerated erosion processes. Degradation due to 

salinity and sodicity varies geographically with climate, agricultural activities, irrigation methods and land 

management policies and is mainly restricted to irrigated farming systems. Causative factors are of intrinsic 

origin, seawater intrusion or irrigation from groundwater with elevated salt content. Degradation due to 

contamination is mainly found in countries with high population, high oil production or heavy mining. In 

irrigated farming systems with overuse of chemicals, the load of toxic elements in groundwater is increased. 

Salinity has greatly reduced crop yields and increased economic annual losses across the region to nearly 

US$1 billion, equivalent to as much as US$1 604 ha -1 to US$2 748 ha -1 . In some countries the reduction in soil 

productivity was estimated to be in the range of 30-35 percent of the potential productivity. 

Responses to degradation caused by erosion include improving soil resilience by increasing C inputs. This can 

be achieved using organic manures, compost and synthetic soil conditioners and soil conservation measures 

on sloping lands. Policies and regulation and socio-economic factors at individual country level were found 

to help reverse land degradation due to erosion. Ways of reclaiming salt-affected soils include: salt leaching 

and drainage interventions, crop-based management, chemical and organic amendments, fertilizers, salt 

tolerant plants, crop management and phytoremediation. Measures to contain degradation caused by oil 

contamination include farming techniques that partly eliminate hydrocarbons through decomposition, 

and bio-remediation using some grass species. With effective desertification control, the potential annual 

C sequestration rate could reach values between 0.2 to 0.4 Pg C yr -1 , compared to the 1.0 Pg C yr -1 in drylands 

worldwide. Ranking of soil threats is given in Table 13.3. 


430 


Threat to soil 

function 


Salinization 


Organic 

carbon change 

Contamination 

Sealing 

Compaction 

Loss of soil 

biodiversity 

Summary 

Wind erosion and dust 

storms are a problem 

throughout the region. 

Sand stabilization in 

source areas is difficult and 

expensive to undertake. 

Water erosion can be 

controlled with adaptive 

management. 

Salinization is a widespread 

problem in the region due 

to the high temperatures, 

inappropriate irrigation 

practices and sea water 

intrusion in coastal 

areas. There is adequate 

research and technical 

knowledge in the region to 

counteract the problem. 

Socio-economic conditions 

hamper widespread 

implementation in some 

countries. 

High temperatures 

throughout most of the 

region result in a very 

high turnover of soil 

organic Carbon. SOC 

change is sensitive to soil 

management changes. 

Contamination is locally 

a significant problem in 

the region particularly 

in urbanized areas that 

produce waste dumped on 

the land 

and in oil producing areas. 

Substantial expansion of 

housing, quarrying and 

infrastructures is a 

concern. There are no 

reliable data on sealing and 

land take. 

Compaction is a problem 

where heavy clay soils are 

intensively tilled 

(e.g. rainfed and irrigated 

Vertisols) and to a lesser 

extent is caused 

by off-road vehicles. 

The extent of loss of soil 

biodiversity due to human 

impact is largely unknown 

in the NENA region. 

More studies need to be 

undertaken to understand 

the scope of the problem. 



Confidence 

In 

In trend 

condition 


431 



acidification 

Nutrient 

imbalance 

Waterlogging 

Given the dry conditions 

throughout most of the 

region, acidification 

is restricted to some 

coastal areas with higher 

rainfall. 

Nutrient imbalances occur 

in areas with continuous 

cultivation 

where nutrients are lost 

in harvested crops and no 

engagement in fallowing, 

manuring or mineral 

fertilizer application. 

Waterlogging is a very 

localized problem in the 

region limited to 

flash floods, heavily 

irrigated areas and 

excessive rise in subsoil 

water level. 

Table 13.3 Summary of soil threats: Status, trends and uncertainties in the Near East and North Africa 


432 


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of Soil Changes in North America 

Regional Coordinator: Hempel, J. (ITPS/United States) 

Regional Lead Author: D. Pennock (ITPS/Canada) 

Contributing Authors: Adams, M-B. (United States), Basiliko, N. (Canada), Bedard-Haughn, A. (Canada), 

Bock, M. (Canada), Cerkowniak, D. (Canada), Cruse, R. (United States), Dabney, S. (United States), Daniels, 

L. (United States), Drury, C. (Canada), Fanning, D. (United States), Flanagan, D. (United States), Grayston, S. 

(Canada), Harrison, R. (United States), Hempel, J. (ITPS/United States), Lobb, D. (Canada), Parikh, S. (United 

States), Reid, D.K. (Canada), Sheppard, S. (Canada), Smith, C.A.S. (Canada), Watmough, S. (Canada). 


442 

in North America


Although Canada and the United States of America have a long history of collaborative research activity in 

soil science, there have been no previous attempts at a regional assessment of threats to soil functions. Nor 

are there any ongoing institutional arrangements that coordinate soil assessment or management across the 

two countries, unlike trans-border water issues – which are adjudicated by an International Joint Commission 

– or atmospheric issues, such as when cross-border problems with acidification were the focus of the Air 

Quality Agreement of 1991. Cross-border coordination is further complicated by the lack of a common soil 

classification system between the two countries. 

In the absence of a regional reporting system for threats to soil functions, this chapter draws on the 

appropriate national reporting systems and on the expertise of leading soil scientists where national 

assessments do not exist. 

The main data source used for Canada is the Agri-Environmental Indicators report series, which was 

developed by Agriculture and Agri-Food Canada (AAFC). This series began in 1993 with the intent of producing 

science-based environmental indicators specific to the agriculture and agri-food sector. The work presented 

in this chapter is drawn from the forthcoming 4th Agri-Environmental Indicators report (Clearwater et al., 

2015). The series estimates change in the indicators in five-year periods beginning (for most indicators) in 

1981. Indicators that assess primary agriculture are calculated using mathematical models or formulas that 

integrate information on soil, climate and landscape, mainly derived from the Soil Landscapes of Canada (SLC) 

(Soil Landscapes of Canada Working Group, 2007), with information on crops, land use, land management 

and livestock from the Census of Agriculture and other custom data sets from provincial agencies, the private 

sector, remote sensing, etc. Information on the specific indicators is available in AAFC (2013). 

The 4th Agri-Environmental Indicators report provides information for soil erosion, change in SOC, and 

nutrient imbalance, and this information is featured in Sections 4.3 and 4.4 of this chapter. Leading Canadian 

scientists selected by the Canadian Society of Soil Science provided information on the remaining threats. 

The major information source used for the United States is the National Resources Inventory (USDA, 

2013a). This report provides a range of land use and management statistics and national estimates for sheet 

and rill erosion and for wind erosion. The report provides data for the period 1982-2010. Data are gathered 

annually by the National Resources Conservation Agency and major reports are released at five-year intervals. 

Information on specific threats such as salinization was also provided by the STATSGO 2 

database (Soil Survey 

Staff, 2014). Leading United States soil scientists selected by the Soil Science Society of America also provided 

information for the United States. 

14.2 | Regional stratification and soil threats 

14.2.1 | Regional stratification and land cover 

The spatial framework used in this chapter is the multi-level Ecological Regions of North America developed 

by the Commission for Environmental Cooperation (Commission for Environmental Cooperation, 1997). The 

Level II ecoregions are used as a consistent geographical reference (Figure 14.1). 

The contiguous 48 United States, Hawaii, Puerto Rico, and the United States Virgin Islands cover almost 

88 Bha of land and water. About 71 percent of this area is non-Federal rural land – nearly 57 billion ha (USDA, 

2013a). The non-Federal rural lands of the United States are predominantly rangeland (165 million ha), forest 

land (166 million ha), and cropland (146 million ha) with smaller areas composed of developed land, pastureland 

and water. 


443 

in North America

Rangeland is dominant in the western half of the United States in the Cold Deserts, Warm Deserts, Great 

Plains, South Central Semi-Arid Prairies and Western Cordillera ecoregions. Forest is the dominant land cover 

in the northwest, north central and eastern one-third of the country, in the Western Cordillera, Mixed Wood 

Plains, Ozark, Ouachita-Appalachian Forests ecoregions. 

Cropland is the dominant land cover in the Central United States Plains, South Eastern United States Plains, 

the Mississippi Alluvial and Southeast United States coastal plains, Temperate Prairies, West Central Semi- 

Arid Prairies, South Central Semi-Arid Prairies and Mediterranean California ecoregions. 

Cropland in the United States increased by about 1 million ha from 2007 to 2010, following a steady decline 

in area in the previous 25 years. These gains can be attributed to land withdrawn from the Conservation 

Reserve Program as grain prices reached near-record levels. This led to an increased threat to the soil resource 

from erosion and loss of terrestrial C. The National Resources Inventory tracks the cropland area for specific 

conservation measures such as terracing. However, according to the Conservation Technology Information 

Center there has not been a national survey of crop residue management practices since 2004; at that time 

45.6 million ha was in conservation tillage out of a total of 112 million ha of cropland (Conservation Technology 

Information Center, 2014). 

Total farmland in Canada increased from 1981 (65.9 million ha) to 2006 (67.6 million ha) (all values from 

AAFC, 2013). The largest agricultural region (54.7 million ha) occurs in the Temperate Prairies and West-Central 

Semi-Arid Prairies ecoregions in southern Alberta, Saskatchewan, and Manitoba. In 2006 approximately 29 

million ha of farmland in this region was cropped, primarily to cereal grains, oilseeds and pulse crops, and 18 

million ha was in pasture. Approximately 1.4 million ha was in tillage summer fallow, where land is fallowed 

during the growing season with weed suppression by one or more tillage operations. The area under tillage 

summer fallow has declined greatly from 5.3 million ha in 1991 and this decline has reduced soil degradation in 

this region. 

Farmland in Ontario and Quebec (8.9 million ha) is concentrated in the Mixed Wood Plains ecozone. In 2006 

5.6 million ha of this was cropped, primarily to forages, maize, cereal grains, and oilseeds (soybeans). The area 

of pastures in this region has declined from 1.7 million ha in 1991 to 1.1 million ha in 2006. Tillage practices in 

this region have also undergone a major shift, with the percentage of cropland under conventional tillage 

decreasing from 80 percent in 1991 to 50 percent in 2006, and conservation tillage and no-tillage increasing 

from 16 percent to 26 percent and from 4 percent to 24 percent respectively over the same period. 

Farmland in British Columbia (2.8 million ha in 2006) is dominated by forage production and pasture 

dispersed through the Western Cordillera and Cold Desert ecoregions. Finally the Atlantic Highlands and 

Mixed Wood Plains ecoregions in Atlantic Canada have 1.1 million ha of farmland, dominantly in forages but 

with significant areas of potato production in New Brunswick and Prince Edward Island. 

Overall the greatest shift in Canadian agriculture has been the adoption of no-till and reduced tillage 

systems. The 2011 Census of Agriculture (Statistics Canada, 2015) reports that of 29.6 million ha seeded in 2011, 

16.7 million ha were in no-till and a further 7.2 million ha were in tillage systems that left most residue on the 

soil surface; only 5.6 million ha were in conventional tillage e.g. with most residue turned into the soil. 

Wetlands are extensive in both Canada and the United States and are subject to considerable alteration by 

human activity. Bridgham et al. (2006) estimate that the historical area of wetlands in Canada of 150 million 

ha has been reduced to 130 million ha by land-use conversion. In the contiguous United States, the same study 

estimated a reduction from 90 million ha historically to current levels of 43 million ha. 


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The area of forested land in Canada is estimated at 348 million ha (Natural Resources Canada, 2012). Forest 

is the dominant land cover in the Northern Forests, Boreal Plain, Boreal Cordillera, Western Cordillera and 

Marine West Coast Forest ecoregions. Forest harvest activity has decreased in the period from 2004, and 

approximately 0.7 million ha were harvested in 2010 (Natural Resources Canada, 2012). Canada’s National 

Forest Inventory shows that gross deforestation rates were typically in the order of 64 400 ha yr -1 circa 1990, 

and decreased to 44 800 ha yr -1 by 2010, corresponding to about 0.02 percent of Canada’s forest area (Natural 

Resources Canada, 2012). Deforestation minus afforestation (according to UNFCCC definitions) amounts to 

a net loss of 35 000 ha yr -1 . Overall, there is a definite decrease in total deforestation rate from the 1990s to 

present, which is expected to continue in coming years, but at a lower rate of decrease (Masek et al., 2011). 

The expansion of agriculture is the largest source of forest conversion, accounting for two thirds of gross 

deforestation. Urban and industrial development is the next largest driver at approximately 17 percent, 

followed by forestry at half that rate. The Boreal Plains ecozone spanning central Alberta, Saskatchewan, 

and Manitoba (generally termed the Prairie provinces) was the dominant location of deforestation over the 

1990–2008 time period, contributing just under half the nation’s deforestation for most years, largely due to 

agricultural conversion (Masek et al., 2011). 

The Tundra (233 million ha), Taiga (200 million ha) and Arctic Cordillera (21.7 million ha) ecoregions are 

vulnerable to effects of climate change. Permafrost soils in these regions are estimated to contain 39 percent 

of all organic C in Canada (Tarnocai and Bockheim 2011) and hence the interaction of soils and climate in these 

regions is a major concern. 


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Figure 14.1 Level II Ecological regions of North America. Source: Commission for Environmental Cooperation, 1997. 


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14.3 | Soil threats 

This section focuses on the status and trends for six threats to soil functioning: acidification, contamination, 

salinization, sealing/capping, compaction, and waterlogging. Four threats that are judged to be the most 

serious (erosion, nutrient imbalance, carbon change, and soil biodiversity) are discussed in greater detail in 

the following section. 


While many native forest communities are well-adapted to strongly acidic (pH < 5.5) soil conditions, most 

managed agricultural and horticultural plants suffer from enhanced metal phytotoxicity (Al, Fe, Mn, etc.), 

reduced N and P availability, and decreased microbiological activity because soils have become acidified 

below their optimum range. Excessive soil acidification also poses environmental risks of enhanced surface 

water acidification, sediment losses due to loss of vegetation, and increased loadings of soluble metals into 

groundwater. Soil acidification is enhanced by a range of anthropogenic effects, including excessive inputs of 

acidic atmospheric deposition, intensive removal of aboveground biomass, and exposure of sulfidic materials 

by mining, construction, dredging, and other disturbances. 

Acidic deposition from rain that has low pH and significant amounts of sulphate and nitrate contributes 

to base cation depletion and soil acidification in industrialized regions of the world (Meinz and Seip, 2004). 

These effects, however, have been documented only rarely in the United States. Coarse-textured soils in highaltitude 

forests that were originally low in pH and base saturation are particularly susceptible to degradation 

and loss of function. Finer-textured and more highly buffered soils are much more resistant to the negative 

effects of acidic deposition. The Clean Air Act Amendments of 1990 resulted in significant declines in sulphate 

emissions (http://nadp.sws.uiuc.edu/ntn/), although emissions of nitrogen oxides have remained elevated due to 

transportation and agricultural sources. Some lakes in the Adirondacks have shown significant improvement 

in water quality as a result of the Clean Air Act Amendments. Research by Driscoll et al. (2001) has shown, 

however, that tighter controls on atmospheric emissions will be needed if soil and stream chemistry in this 

region is to return to pre-industrial levels in a reasonable time range. Alternatively, aerial application of calcium 

sources such as wollastonite (CaSiO 3 

) to acidifed catchments can accelerate the return of base saturation to 

pre-industrial levels (Johnson et al., 2014). However, the widespread applicability of this reclamation approach 

is limited. Localized, highly acidic deposition from heavy metal smelter and other industrial facilities also 

impacted large areas of land such as the Copper Basin in Tennessee, where many square miles of land were 

denuded by open air smelting of metal sulphide ores. Current concerns centre on the effects on soil fertility 

and acidification from the interaction between intensive biomass harvesting and acidic deposition on forest 

soils (Adams et al., 2000). 

Sulfidic materials (as defined by Soil Taxonomy (Soil Survey Staff, 1999)) are routinely exposed by mining, 

construction, and dredging activities. This exposure can rapidly lower the pH of local soils and water to < 4.0 

via sulfuricization processes (Kittrick, Fanning and Hossner, 1982). While most active mining operations now 

isolate sulfidic materials from contact with groundwater and surface water via the application of appropriate 

acid-base accounting procedures (Skousen et al., 2002), construction-related impacts have become 

increasingly common since the 1970s due to larger scale excavations and the construction industry’s lack of 

recognition of risk (Fanning et al., 2004; Orndorff and Daniels, 2004). Once exposed, sulfidic materials require 

large inputs of liming agents (e.g. ~ 31 Mg of agricultural lime per 1000 Mg material for 1 percent pyritic-S) to 

become properly neutralized and stabilized. 

In Canada, the major risks associated with soil acidification occur in forested areas. As in the United States, 

areas most at risk from acidification are in regions dominated by coarse-textured soils that have low base 

cation weathering rates and that receive high levels of acid deposition (Ouimet at al., 2006). These areas 


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of coarse-textured soils include much of the Softwood Shield and Mixed Wood Shield ecoregions and the 

southern, coastal parts of British Columbia (Aherne and Posch, 2013) in regions where glacial parent materials 

were derived from igneous rocks. Both the loss of essential base cations and the mobilization of metals such as 

Al and Mn can have adverse impacts on forest vegetation. It has been proposed that the ratio of base cations or 

Ca to Al (e.g. base cations (BC/Al or Ca/Al) in soil solution is a useful indicator of the potential risk to trees from 

soil acidification (Cronan and Grigal, 1995; Sverdrup and Warfvinge, 1993). Critical loads are also increasingly 

used to estimate the risk of soil acidification (Whitfield et al., 2010). 

Aherne and Posch (2013) estimated critical loads for acid deposition for upland forest soils in Canada (~2 600 

000 km 2 ). They reported that in 2006, because of acid deposition levels, 4.5 percent of the mapped area (~100 

000 km 2 ) was at risk based on a BC/Al ratio of 1.0; and 20.3 percent (~500 000 km 2 ) of the area was at risk 

based on a BC/Al ratio of 10.0. Exceedance of the critical load was primarily driven by elevated anthropogenic 

emissions from large point sources, such as the activities in the Athabasca Oil Sands region and in ore smelting 

near Sudbury, Ontario. In addition, exceedance in central and eastern Canada was associated with long-range 

(transboundary) air pollution and emissions from shipping along the St. Lawrence River. 

In Canada, national emissions of SO 2 

and NOx decreased by 21 percent and 3 percent, respectively, between 

2008 and 2010 (Canadian Council of Ministers of the Environment (CCME), 2103). This decrease reduces the 

risk of soil acidification in Canada. Emission reductions, however, vary by province. The bulk of the reduction 

in SO 2 

and NOx emissions has been in Ontario in central Canada. Minimal decreases or even increases in 

emissions, associated primarily with the oil and gas industry (CCME, 2013), have occurred in British Columbia 

and Alberta in western Canada. Currently, the risk of soil acidification caused by acid deposition is generally 

decreasing over much of eastern Canada but is unchanged or increasing in parts of western provinces, such as 

Alberta and British Columbia. 


Soils can be compromised via industrial, mining, municipal, residential and agricultural activities. In North 

America, metals (Pb, Cd, Cr and As), salts (Na and K), pesticides (herbicides and insecticides), pathogens 

and nutrients (N and P) contaminate soils to varying degrees and with great spatial variation. There are also 

chemicals of emerging concern, including engineered nanoparticles, pharmaceuticals and personal care 

products. Perfluorinated compounds are also of concern: they occur in small concentrations but, because of 

their high reactivity or potential to be endocrine disrupting, they may pose significant risks to human health 

and the environment. 

In the United States, there are thousands of organic- and metal-contaminated sites of varied scope 

and significance. To address, monitor and remediate the myriad of sites, the United States Environmental 

Protection Agency oversees the Superfund programme, which is charged with the clean-up of the nation’s 

hazardous waste sites. This effort follows the National Priorities List (NPL), which defines the known releases 

or threatened releases of contaminants in the United States and its territories (EPA, 2014a). As of 29 September 

2014, there are 1 322 final sites on the NPL with 1 163 having completed measures to address the contamination 

threat and an additional 49 proposed sites (Figure 14.2). In addition, there are vast areas of low-level soil 

contamination across the United States which are not monitored by the EPA. 


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Figure 14.2 Map of Superfund sites in the contiguous United States Yellow indicates final EPA National Priorities List sites and red 

indicates proposed sites. Source: EPA, 2014a. 

In Canada, because it has a huge expanse of soil and a relatively small population, soil contamination in 

a spatial context is a relatively minor issue. However, the most agriculturally productive soils and greatest 

density of population and industry occur concomitantly along the narrow region close to the southern border. 

This is also the region where there is the greatest potential for soil contamination. In addition, the hinterland 

has widely dispersed petroleum and mineral resource industries that form hot spots of soil contamination 

(Doyle et al., 2003). 

More insidious is non‐ point‐ source, dispersed contamination (Chan et al., 1986). For example, field crop 

soils surveyed throughout the Mixed Wood Plains ecoregions of southern Ontario in Canada showed elevated 

levels of Ba, Cd, Mo, Pb, Sb, Se, Nb, U and Zn, which were speculatively attributed to non‐ specific urban sources 

such as road dust (Sheppard et al., 2009). Watmough and Hutchinson (2004) came to a similar conclusion 

about Pb in forest soils of Southern Ontario. Toxicity in soil from such sources, however, is a relatively remote 

possibility. 

There is concern about soil contamination by agricultural activities, especially as farms increase in size 

and effectively become industrial point sources. For example, soils in areas of livestock facilities have been 

found to have metal levels that exceed Canadian soil quality guidelines (Sheppard and Sanipelli, 2012). Some of 

these metals came from livestock pharmaceuticals (e.g. Bi in teat dips). The contribution of livestock manures 

containing antibiotic residues to the development of antibiotic-resistant genes in the environment is a growing 

public concern. In a few cases, the naturally occurring, trace element bioavailability of some Canadian soils 

has resulted in food crops with amount of elements that exceed guideline concentrations. The most notable 

are spatially isolated cases of Cd in durum wheat and sunflowers (Grant et al., 1998). 

As industrialization and urbanization increase, concomitant with increased agricultural activities in 

decreasing land areas, the potential for soil contamination remains an important issue. Although soil 

contaminants in the United States and Canada are ubiquitous in areas close to human populations, the 


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specific threats posed by these contaminants to human health and environmental quality are not well defined. 

There is a need for improvements in assessments of soil contamination to better protect human health and 

environmental quality and ensure food safety and security. 

14.3.3 | Soil salinization 

Soil salinization is a serious threat to the ecosystem services provided by the soil resource with regard to 

food and fibre production in many parts of North America. The movement and accumulation of salts that 

cause saline conditions in the soil are affected by the soil water balance. Processes such as climate shifts, 

improper irrigation and drainage, farming and management practices affect this balance. Soil salinity is a 

dynamic soil condition and can spread or become more severe in areas that are already saline, especially if the 

land is not managed properly. 

These concerns are especially prevalent in the western portions of the United States (Figure 14.3) (Soil Survey 

Staff, 2014). Similar threats to food and fibre production are also associated with sodic conditions in the soil. In 

the United States, saline soils occupy approximately 2.2 million ha of cropland and another 31 million ha are at 

risk of becoming saline (USDA, 2011). 

In Canada, a Risk of Soil Salinization (RSS) Indicator has been developed as part of the Agri-Environmental 

Indicators programme to assess the state and trend of the risk of dryland soil salinization on the Canadian 

Prairies as a result of changing land use and management practices. Two of the primary conditions required 

for dryland salinization - water deficits and an inherent salt content in the soil and/or groundwater - occur to 

a significant extent only in the Prairie region of Canada. The risk of salinization on other agricultural lands in 

Canada is negligible. The risk of soil sodicity is not assessed as part of this index and is not believed to be a major 

risk in western Canada. The RSS is derived by calculating a unit-less Salinity Risk Index (SRI) which considers 

a combination of factors that control or influence the salinization process (Wiebe, Eilers and Brierley, 2010). 

In terms of the state of soil salinity in Canada, approximately 1 M ha of surface soils on the Prairies are 

affected by moderate to severe soil salinity (Wiebe, Eilers and Brierley, 2006). In 2011, 85 percent of the land 

area in the agricultural region of the Canadian Prairies was rated as having a very low risk of salinization 

(Figure 14.4). 

From 1981 to 2011, the trend has been a 19 percent increase in the land area in the Very Low and Low 

risk classes. Over the same 30-year period, the land area in the Moderate, High and Very High risk classes 

decreased from 15 percent to 8 percent. These improvements were largely attributed to the reduction in tillage 

summer fallow, mentioned above, and to a 4.8 million ha increase of permanent cover (a 14 percent increase 

from 1981 to 2011). A reduction in risk has been observed in all Prairie provinces. The greatest decline was 

recorded in Saskatchewan, where the area of summer fallow decreased by more than 5 million ha and the 

area of permanent cover increased by more than 3 million ha. Changes in land use and management practices 

have reduced the risk of salinization and indicate a trend towards improved soil health and agri-environmental 

sustainability. 


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Figure 14.3 Areas in United States threatened by salinization and sodification. Source: NRCS 1 

1 http://www.nrcs.usda.gov/wps/portal/nrcs/site/national/home/ 


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Figure 14.4 Risk of soil salinization in Canada 2011. Source: Clearwater et al., 2015. 

14.3.4 | Soil sealing/capping 

The population of North America is approximately 5 percent of the world population (340 million) and it has 

been growing at a rate of 0.9 percent a year, during the last decade. In the United States, the threat of loss of 

soil due to soil sealing and capping consequent on expansion of settlements and infrastructure is significant 

and has been steadily increasing since 1982, as documented by the USDA Natural Resources Conservation 

Service’s National Resources Inventory Program (NRI) (USDA, 2013a). Between 1982 and 2007, it is estimated 

that 16.5 million ha of land were developed for urban or transportation uses. By 2007, the United States had 

an estimated total of 45 million ha developed into urban uses. Of the newly developed land, 41 percent was 

previously forest land, 27 percent was cropland, 17 percent was pasture, and 13 percent was rangeland. In regard 

to land categories, 35 percent of the land in the United States that was developed into urban uses during the 

period of 1982 to 2007 was classified as prime farmland. Prime farmland is land that has the best combination 

of soil physical and chemical characteristics for producing food, feed, forage, fibre and oilseed crops and has 

the soil quality, growing season, and moisture supply needed to economically produce sustained high yields 

of crops when treated and managed at high levels. Prime farmland is also the most economically viable land 

to develop as it typically has the lowest degree of limitations for conversion into urban development. This 

increases the pressure on land with the best soil. Using the 2007 NRI information as a base, the trajectory of 

prime farmland (cropland portion) conversion and the loss of potential food production is immense over the 

next 25 years. With this rate of loss, the United States will lose the equivalent of 10M metric tonnes of corn in 

2022 due to sealing/capping activities. 


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The drivers for soil sealing in Canada are very similar to those in the United States. The growth of metropolitan 

centers has been particularly rapid since the 1990s in areas surrounding Toronto, Kitchener-Waterloo, Ottawa, 

and Vancouver and, most recently, in areas surrounding urban centers in Saskatchewan and Alberta. Between 

1996 and 2006, urban land increased by more than 10 percent nationally. It now totals nearly 0.13 million ha 

(Statistics Canada, 2009), or 0.25 percent of the total land area in Canada. Given land-cover data extrapolated 

from a suburban United States city (specifically, San Jose with 59 percent impervious surface by area (Xiao et 

al., 2013) and the fact that suburbs now make up the majority of Canada’s metropolitan population (Gordon 

and Janzen, 2013), it is estimated that more than 1 300 km 2 of soil was capped through urban expansion 

between the 1990s and 2000s. It is noteworthy that this expansion has occurred largely on highly productive 

soils (Francis et al., 2012; Hofmann, Filoso and Schofield, 2005). Major interregional highway construction 

and expansion (e.g. highway twinning) have also been ongoing, particularly in Ontario, British Columbia, and 

Alberta. Total road length (in two lane equivalents) in Canada increased by more than 17 percent between 

1990 and 2009 (Transport Canada, 2012; United States Department of Transportation, 2014), and, assuming 

a conservative average road width of 10 m, roads now cover more than 10 000 km 2 . However, it is important 

to note that a portion of this impervious area is also included in the estimated urban area described above. 

In the past few decades, areas used for agriculture and forest harvest have decreased (Natural Resources 

Canada, 2014; Francis et al., 2012; Hofmann et al., 2005). Thus, it is assumed that sealing as a result of new 

construction for agriculture service and forest industrial roads is relatively minor. The overall existing extent of 

unpaved roads in Canada that serve the needs of these industries is substantial. 

14.3.5 | Soil compaction 

Soil compaction is an acknowledged threat to the ability of the soil resource to provide a wide range of 

essential ecosystem services, including food and fibre production and maintenance of good water quality. 

Compaction decreases the water infiltration capacity of the soil, increases runoff and erosion, reduces plant 

growth, and reduces the penetration, size, and distribution of roots. It restricts water and air movement in 

the soil. It also causes nutrient stresses and slow seedling emergence. Soils in North America that are most 

susceptible to the threat of compaction are located in managed agricultural and forested regions. 

The most common cause of soil compaction (or the formation of hardpans) is agricultural traffic, including 

tractors, harvesting equipment and implement wheels on moist soils where soil moisture is at or above field 

capacity. Wheeled traffic compacts soil aggregates, in some cases destroying the aggregates completely, 

which results in a dense soil with few large pores and limited aeration. Compaction can reduce yields up to 

50 percent in some areas, depending upon the depth of compaction and its severity (Wolkowski and Lowery, 

2008). 

The type and condition of a soil have an effect on the potential of compaction to occur. Soils low in organic 

matter tend to be more susceptible to compaction because their ability to form strong aggregates is decreased. 

Soils high in clay content compact more easily because clay particles adhere to water, which makes it easier 

for them to move against each other. In the United States, Ultisols in the Coastal Southeast United States 

Coastal Plains ecozone are especially susceptible to compaction due to their highly weathered state and low 

organic carbon levels (Simoes et al., 2009). 

In Canada, there has not been a national-scale assessment of compaction since McBride, Joosse and Wall, 

(2000). Generally, fine-textured soils in the Mixed Wood Plains ecozone under cropping systems with a high 

potential for soil structure degradation and compaction (e.g. those for the production of maize, soybeans, 

vegetable and root crops) were judged to have the highest risk of compaction. In the period from 1981 to 1996, 

the area of fine-textured soils under high-risk cropping practices grew substantially (e.g. by as much as 61 

percent in Ontario). This area expansion of potentially compaction-inducing cropping practices has continued 


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to the present day. Agricultural soils in western Canada and the United States are generally believed to have 

a lower risk for soil compaction. The exception is irrigated soil in this region, where the wetter soil conditions 

can contribute to higher compaction levels. In addition, adoption of conservation tillage, which limits 

compaction, has also been lower on irrigated land. 

Soil compaction also occurs in forestry operations throughout Canada and the United States, especially 

where roads and landings have been constructed. Generally, finer-textured soils are at the greatest risk, 

but the amount of organic matter in coarser textured soils can form a strong negative relationship with the 

potential for compaction (Krzic et al., 2004). Compaction has been shown to reduce regeneration of species 

such as aspen (Populus tremuloides Michx.) and white spruce (Picea glauca [Moench] Voss) (Kabzems, 2012), 

but little is known about its general impact on soil functions. 

14.3.6 | Waterlogging and wetlands 

The threat of waterlogging to the sustainable use of soil resources is difficult to assess because waterlogging 

is not considered a threat in all cases. Wetland areas that are nearly or permanently waterlogged provide many 

positive benefits to the environment. Wetlands are some of the most biologically diverse habitats on earth and 

are of great benefit to many species of wildlife. They also act as a filter, trap sediments, improve water quality, 

are a carbon sink, and reduce peaks of floodwater runoff. Until 2014, the Wetland Reserve Program in the 

United States was a voluntary programme offered to landowners to protect, restore, and enhance wetlands 

on their property. Nearly 1 million ha has been enrolled in this programme (USDA, 2014). 

Much of the drainage of wetlands in North America has been concentrated on freshwater mineral wetlands 

(Bridgham et al., 2006), which are wetlands dominated by Gleysolic soils rather than organic soils. Bridgham 

et al. (2006) estimate that both the United States and Canada have experienced over 50 percent conversion of 

this class of wetlands (e.g. a reduction from 36 million ha to 16 million ha in Canada and from 76 million ha to 

31 million ha in the United States). 

In North America, the primary driver of large-scale waterlogging on non-wetland soils is flooding due to 

dam construction for hydroelectric power, to flood control measures and to mining activities (Maynard et al., 

2014). This includes both upstream flooding associated with dam construction and high spring water levels 

and downstream flooding associated with controlled releases during high-water periods and for hydroelectric 

power generation during the winter. 

Another indirect driver is deforestation, which can reduce infiltration and/or evapotranspiration. It has 

been proposed that reduced evapotranspiration in northern Alberta contributes to a significant rise in the 

water table when deforestation is coupled with wet climatic periods (Carrera-Hernandez et al., 2011). 

Concern about waterlogging has been growing due to the substantial increase in the frequency and severity 

of extreme precipitation events in recent years in some regions (Brimelow et al., 2014), even though this may 

not indicate an overall increase in mean annual precipitation for most regions. Waterlogging related to all of 

the above causes is exacerbated by increased precipitation. 

14.4 | Major soil threats 

Four threats to soil functions were selected as major threats and are covered in more detail in this section. 

The main criterion for their selection was the area of land affected by these threats – all four of these threats 

operate in most agricultural (and many non-agricultural) landscapes, whereas the threats covered in Section 

14.3 tend to be more locally focused. Data on these four threats for Canada are covered in more detail in the 

case study in the following section (14.5). The present section focuses on North American-scale drivers and 

specific results for the United States of America. 


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Soil erosion in the United States and Canada accelerated after the arrival of European settlers, who cleared 

extensive areas for agriculture and subsequently ploughed and overgrazed the land (Montgomery, 2008). 

Soils rapidly degraded and erosion increased as settlement spread from east to west. In the United States, 

erosion was greatest on the east coast in the early 1800s, in the mid-south during the early 1900s, and in the 

Great Plains during the ‘Dust Bowl’ era in the 1930s. Some badly degraded lands were abandoned and then 

reverted to secondary growth forests, a process that slowed erosion rates. Wind erosion was very significant 

in Prairie Canada during the 1930s. Soils that were badly degraded due to wind erosion were subsequently 

stabilized and converted to permanent pasture. 

Agricultural mechanization, commercial nitrogen availability, and federal policies encouraging maximum 

crop production led to cash-crop intensification throughout the middle of the 20th century in both the United 

States and Canada. Forage-based rotations were shortened or eliminated, field sizes were increased by the 

removal of hedgerows and fences, and tillage intensity remained high. As a result, the potential for soil erosion 

increased during this period. In the late 20th and early 21 st centuries higher yielding varieties and improved 

herbicide technology supporting the adoption of conservation tillage helped reduce the potential for water 

and wind erosion. Federal farm programmes in the late 20th and early 21 st centuries had both favourable and 

unfavourable impacts on soil erosion rates. In Canada, the most significant cropping change was the major 

reduction in summer fallow (e.g. the practice of leaving land fallow for one growing season and suppressing 

weed growth by one or more tillage events) in the two Prairie ecoregions in Canada. This change substantially 

reduced the risk of wind erosion. 

In the United States, the National Resources Inventory (NRI) has reported statistically robust estimates of 

water (sheet and rill) erosion and wind erosion on privately owned cropland, since 1982 at fiveyear intervals 

(USDA, 2013a), based on a wide monitoring network and an assumed historic average climate for each 

location. The estimated decrease in sheet and rill erosion between 1982 and 2002 was 39 percent, and that 

between 1982 and 2010 was 41 percent. In the same periods, wind erosion decreased by 41 percent and 46 

percent, respectively. 

In 2010, the most intense sheet and rill erosion was in the Temperate Prairie and Mixed Wood Plains 

ecoregions of the Midwest United States, in the adjacent area of the South-eastern United States Plains 

ecoregion and in the Palouse (Cold Desert ecoregions in the state of Washington). These areas and the West- 

Central Semi-Arid Prairies ecoregion also had the highest wind erosion rates. 

If 2010 NRI estimates of sheet and rill erosion plus wind erosion are averaged across all United States 

cropland, the average annual rate is about 10 Mg ha-1, and 57 percent of this is due to sheet and rill erosion 

(USDA, 2013a). Tolerable annual soil erosion rates (‘T’) used in the United States typically range from 2 to 11 

Mg ha-1, depending upon the soil type. The criteria used to calculate T have been widely criticized (Johnson, 

1987) and the overall average soil loss rate is one order of magnitude greater than estimated soil renewal rates, 

which are less than 1 Mg ha -1 peryear (Alexander, 1988; Montgomery, 2007). In addition, erosion varies spatially 

and may greatly exceed T at any specific NRI point, due to the soil, slope, management and climate at that 

location. 

The state of soil erosion in Canada and the drivers of change are covered in more detail in Section 14.5 of this 

chapter. Soil erosion on undisturbed forested land in Canada is generally believed to be low (Maynard et al., 

2014) and no national estimates exist for rates of erosion on disturbed forest land. 


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Not all forms of erosion are considered in the national estimates and this leads to considerable overall 

uncertainty in estimates. In the United States, the NRI erosion assessment does not include ephemeral 

gully or tillage erosion. Tillage erosion is within-field soil redistribution by tillage implements, which has 

been extensively documented in both the United States and Canada. Both of these processes may result in 

degradation rates comparable to those from wind and water erosion. The best estimate of soil erosion rates 

must include estimates of these processes. When these two processes are included, average United States 

cropland soil erosion rates exceed published soil development rates by more than one order of magnitude. 

In the future, an increased frequency of extreme rainfall events due to climate change will likely increase the 

water soil erosion threat in many parts of the United States and Canada. 

In Canada, the national Agri-Environmental Indicators programme does include an assessment of 

tillage erosion, which is known to be of equal or greater significance than wind and water erosion on some 

landscapes. Gully erosion is not included in the Canadian monitoring system, but its incidence in Canada is 

believed to be limited. Although accelerated erosion associated with forest harvest is a concern, there are no 

recent national-level surveys of its incidence in Canada. 


Many regions of North America have experienced and continue to experience nutrient applications in 

excess of plant requirements. These surpluses lead to elevated levels of N and P in soils, which cause a range of 

environmental problems and are a source of considerable societal concern throughout North America. 

The greatest issue with nutrient imbalance in North America is the impact of elevated N and P levels in 

soil from past and present agricultural activities on water quality. The linkage of elevated soil N and P levels 

to water quality problems ranges from algal blooms due to eutrophication in Lake Winnipeg in Manitoba 

(Schindler, Hecky and McCullough, 2012), at the northern edge of the agricultural zone, to the seasonal 

hypoxia in the shallow coastal waters of the Louisiana shelf in the northern Gulf of Mexico, at the southern 

end of the agricultural zone (Alexander et al., 2008). 

Estimates on excess nutrient levels presented by Foley et al. (2010, Supplementary Information Maps S6e 

and S6f) show that excess application of N continues in many regions of North America whereas little excess 

application of P occurs. Excess N application of between 60 to 100 kg ha -1 occurs in much of the Temperate 

Prairie and Mixed Wood Plains ecoregions in both the United States and Canada, throughout the Mississippi 

River valley, and in pockets in the Southeast United States Coastal Plains ecoregion. Hence over-application of 

N is an on-going issue whereas elevated P levels may be largely due to historical over-application. 

The linkage between agricultural practices and N and P loads in waterways has been shown by many 

studies. For example, recent studies on the Missouri River (Brown, Sprague and Dupree 2011) and the entire 

Mississippi River basin (Alexander et al., 2008) using the Spatially Referenced Regressions On Watershed 

Attributes (SPARROW) model by United States Geological Survey (2014) researchers found that the majority 

of both total N and total P in these waterways is from agricultural land. Specifically, 52 percent of total N 

reaching the Gulf of Mexico was from maize-and soybean-producing land with a further 14 percent from all 

other crops in the basin. Some 37 percent of total P was from rangeland/pasture land with a further 25 percent 

from maize- and soybean-producing land. Considerable regional variations occur. For example, in the western 

sections of the Missouri River basin where cattle grazing is the dominant land use, as much as 34 percent of 

total N was from manure whereas in the Mississippi River basin as a whole, only 5 percent of total N was from 

rangeland and pasture land. 


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The relationship between soil properties, management, and particular N and P fractions is complex 

(Sharpley and Wang, 2014; Harmel et al., 2006). Harmel et al. (2006) examined the relationship between soil 

and site attributes and N and P fractions in nutrient loads from watersheds in 15 United States states and two 

provinces of Canada. Particulate N and P loss contributed, on average, three times as much as dissolved forms 

to loads, indicating the overriding effect of soil erosion and transport on N and P loads. Median particulate 

N loads were greater in areas of conventional tillage (which experience higher erosion rates on average) and 

lower in areas of conservation tillage and no-till land, although no differences were observed for particulate 

P. There was a weak relationship between soil test P and all forms of P load, again indicating the importance 

of existing or legacy soil P content. Dissolved N and P loads were highest in areas of no-till land. The build-up 

of P at the soil surface in no-till systems was also implicated in the increase since 1995 of dissolved P load in 

the Maumee River system (which drains into Lake Erie), although the response of this system to management 

changes was very complex (Sharpley and Wang, 2014). 

Excess soil nitrogen can also leach from the soil as nitrate. This threat is most severe in situations where 

shallow aquifers are used as potable water sources in humid or sub-humid climates with coarse-textured 

soils utilized for intensive agricultural production. Examples in Canada include Prince Edward Island, the 

Abbotsford aquifer, BC and Kings County, Nova Scotia. The indicator of risk of water contamination by 

nitrogen (IROWC-N) in the Canadian Agri-Environmental Indicators doubled from ~5 mg N L-1 in 1981 to a value 

approaching the Canadian drinking water guideline for nitrate (10 mg N L-1) in 2006 in the Atlantic Highlands 

and in the Canadian portions of the Mixed Wood Highlands (AAFC, 2013). 

Country-specific information for nutrient imbalance in Canada is given in Section 14.5 of this report. 

N levels in excess of plant requirements in soils are also linked to other environmental issues, especially 

the enhanced release of the potent greenhouse gas, N 2 

O, from soils. In both Canada and the United States, 

agriculture accounts for 6 to 7 percent of total GHG emissions (EPA, 2014b; Environment Canada, 2013b). 

Emissions of N 2 

O from agricultural soils account for 75 percent of the agricultural total in the United States and 

65 percent in Canada. The highest N 2 

O emissions occur under anaerobic conditions and hence are intimately 

linked to changes in waterlogging in agricultural landscapes. Health concerns are also linked to forms of N in 

groundwater and fertilizer application, although the direct link to human health can be difficult to ascertain 

(Manassaram, Backer and Moll, 2006). 


Assessment of soil organic carbon (SOC) in the United States and Canada currently includes a number of 

approaches aimed at either directly measuring or modelling SOC change over time. Both countries utilize 

national-scale modelling of SOC change for GHG emissions inventories and reporting. 

Models of SOC changes currently show increases in the United States. These increases in SOC are primarily 

due to less intensive agriculture (McLauchlan, Hobbie and Post, 2006) and reduced tillage intensity (West 

et al., 2008). Models generally predict that converting areas from native vegetation (e.g. prairies and native 

forest) to cropland and forest plantations results in decreased SOM; however, in some cases, high-productivity 

crops and SOC-conserving management may enhance SOC. For instance, Ogle et al. (2010) estimated that 

SOC in United States croplands increased by 14.6 and 17.5 Tg yr -1 during 1990-1995 and 1995-2000, respectively, 

primarily due to reductions in tillage intensity. 

For agricultural soils in Canada, national-level modelling indicates that improvements in farm management 

have resulted in a dramatic shift from stable SOC levels (e.g. additions are equal to losses) during the mid- 

1980s, to a situation where the majority of cropland had increasing SOC levels in the mid-1990s through to 

2011 (discussed in more detail in Section 14.5 of this chapter). 


457 

in North America

There is currently high uncertainty associated with SOC models (Ogle et al., 2010) such as CENTURY 

(National Resource Ecology Laboratory, 2007), the SOC model most commonly used in the United States and 

Canada to predict SOC change. Improving model parameterization and adding additional SOC measurements 

over time could help reduce uncertainty. This information could be used to better calibrate, estimate and 

potentially reduce model uncertainty as well as to directly track SOC change (Jandl et al., 2014). Nationallevel 

modelling is also limited by the lack of data on SOC stocks and change deeper in the soil profile. Models 

and researchers generally consider subsurface SOC as a relatively stable pool. Little direct evidence has been 

provided, however, that SOC stabilized under previous conditions will remain stable with changing conditions, 

such as with climate change. Some studies have shown that previously stable SOC may be rapidly converted 

to CO 2 

(Fontaine et al., 2007; Fang et al., 2005). 

There are currently two major United States-wide efforts that sample soil over time, the Forest Inventory 

and Analysis (FIA) by Gillespie, 1999, covering United States forests, and the Rapid C assessment (USDA, 

2013b), which includes all vegetated areas. The FIA samples soil to a maximum depth of 20 cm, so its utility 

in monitoring whole-profile SOC over time is limited (Waltman et al., 2010; Jandl et al., 2014). The Rapid 

C assessment (USDA, 2013b) samples soil profiles to 100 cm depth. It is currently uncertain what depth is 

required to truly understand SOC and potential changes in SOC (Harrison, Footen and Strahm, 2011). Some 

studies have shown that results of monitoring SOC change vs. land management depend more on maximum 

soil sampling depth than on treatments (Liebig et al., 2005; Khan et al., 2007; Harrison, Footen and Strahm, 

2011). 

Field data on losses of SOC in Canadian soils after conversion from native land to cropland, and for different 

tillage, crop rotation and fertilizer management practices were compiled from a total of 62 studies by 

VandenBygaart et al. (2003). They demonstrated that 24 ± 6 percent of the SOC was lost after native land was 

converted to agricultural land. In the past two decades, no-till (NT) increased the storage of SOC in Mollisols 

(Chernozems) of the two Prairie ecoregions by 2.9 ± 1.3 Mg ha–1; however, in the moister soils of central and 

eastern Canada, conversion to NT did not increase SOC. More recent studies using meta-analyses (Congreves 

et al., 2014) of long-term agricultural management effects on SOC in Ontario indicate trends towards higher 

SOC with NT than under conventional tillage practises. Crop rotation was found to lead to higher SOC than 

when continuous maize was grown, and the application of N fertilizer led to an increase in SOC compared to 

when no N fertilizer was applied. 

Carbon change in managed forests (232 million ha) in Canada is assessed as part of the National Inventory, 

primarily using the CBM-CFS model (Kurz et al., 2009). Carbon emissions from the dead organic matter and 

soil pools are lumped together. Results showed a small increase in emissions from 2000-2007 due to the 

short-term effect of past disturbances, especially insect infestation. However, the values decreased from 2008 

until 2012 and have returned to long-term levels. Freshwater mineral wetland soils in the Prairie ecoregions 

are also important carbon reservoirs, and approximately 70 percent of these were impacted by agricultural 

activities in 2005 (Bartzen et al., 2010); however there is no regional estimate of associated carbon change. 

Permafrost soils are classified as Cryosols in Canada and cover 2.5 million km 2 ; they are estimated to 

contain 39 percent of all organic carbon in Canadian soils (Tarnocai and Bockheim, 2011). The greatest driver 

of change in these soils is climate change. The IPCC 5th Assessment Report (Clais et al., 2013) states that there 

is high confidence that reductions in permafrost due to warming will cause thawing of some currently frozen 

carbon, but there is low confidence on the magnitude of CO 2 

and CH 4 

emissions to the atmosphere due to the 

complexity of the biogeochemical processes involved. 


458 

in North America

14.4.4 | Soil biodiversity 

Soil biodiversity refers to the myriad of organisms living in the soil, ranging from the smallest microorganisms 

(e.g. bacteria, archaea and fungi) to soil invertebrates. Up until the advent of molecular biology and its use 

in soil science in the 1990s, the assessment of soil biodiversity was done using morphological methods. 

Methods to study soil biodiversity are improving constantly with the application of sequencing technologies, 

complementing the morphological assessments. 

As a result of our paucity of knowledge, assessing threats to soil biodiversity is very difficult. There is no 

reference baseline data for these organisms, nor do scientists have the ability to estimate the true numbers 

of soil organisms, particularly microorganisms. Many organisms have not been described and overall we need 

a better understanding of the biogeography of soil organisms in North America (Nunez and Dickie, 2014). 

Advances, however, have been made. For example, Taylor et al. (2014) completed a comprehensive survey of 

fungi in black spruce (Picea mariana) sites in Alaska and recorded 1 002 taxa in this system. They reported a 

fungus: plant ratio of 17:1. 

However, as soil organisms are so intricately tied to aboveground plant species, threats to plant species 

such as habitat loss are also liable to affect soil organisms (Wardle et al., 2004). Evidence of this exists. For 

example, it has been shown that removal of logging residues from harvested forest sites is one of the major 

threats to forest fungi and insects (Berch, Morris and Malcolm, 2011). Similarly, invasive plant species and 

their mutualistic microbes pose a threat to native mutualist communities (Nunez and Dickie, 2014). Invasive 

alien species are entering North America with increasing frequency due to the growing volume of trade, 

the broadening of trading partners, and the increases in travel and tourism that accompany globalization 

(Environment Canada, 2013a). 

Regional or national-level programmes that monitor soil biodiversity are lacking in North America. In 

Europe, Gardi et al. (2013) used modelling of data from 20 experts to demonstrate that the main pressures 

on soil biodiversity were intensive land exploitation (such as agriculture intensification using tillage, crop 

rotations, and additions of pesticides and herbicides) and changes in land use (such as the reduction in forests 

that have the highest soil biodiversity and the increase in soil sealing) combined with decreasing amounts of 

soil organic matter and increasing numbers of invasive species. 

These modelling results are increasingly being supported by field studies using emerging identification and 

community analysis techniques. Crowther et al. (2014) assessed deforestation effects on soil biodiversity at 

eleven sites in the United States and found that forest removal was generally associated with reductions in 

fungal and bacterial microbial biomass and increases in diversity of taxa. The magnitude of differences due to 

deforestation varied drastically between sites and was best explained by differences in soil texture: the effects 

were greatest in coarse-textured soils and least in fine-textured soils. Crowther et al. (2014) suggest that the 

relationship between soil biodiversity and soil texture offers the potential for mapping regional and national 

patterns of the susceptibility of total (fungal, bacterial, and archaeal) soil biomass to changes in vegetation 

(see Figure 4 in Crowther et al., 2014). Studies based on a meta-analysis of crop rotation found that adding 

one or more crops in rotation increased microbial biomass carbon by 20.7 percent and microbial biomass 

nitrogen by 26.1 percent, indicating the sensitivity of soil microbes to the quantity and biochemistry of crop 

inputs (McDaniel, Tiemann and Grandy, 2014). Other studies, such as those by Wagg et al. (2014), are explicitly 

examining the relationship between decreasing biodiversity in the soil community and changes in ecosystem 

functions, such as carbon sequestration and nitrogen and phosphorus leaching. 

In addition, climate change is considered a threat to soil biodiversity. Unfortunately, there is a scarcity of 

data on this issue. To determine this threat we need to predict soil biodiversity patterns, which is not possible 

with our current knowledge. However, there are indications that soil biodiversity will be reduced by climate 

change. Studies in the Canadian Arctic show that extreme ecosystems contain many unique organisms that 

may become extinct with permafrost melting (Vincent et al., 2009). Wildfire events, which are another threat 

to soil biodiversity, are also predicted to increase because of climate change (Krawchuk et al., 2009). 


459 

in North America

At present, no national or regional assessment on loss of biodiversity can be made for North America, but 

there are signs that such an assessment may be possible in the future. The Global Soil Biodiversity Initiative 

(GSBI) was launched in 2011 and provides a welcome platform for the coordination of research in this area 

(GSBI, 2014). 

14.5 | Case study: Canada 

As discussed in the Introduction (14.1), there are no existing regional maps for threats to soil functions 

for North America. In this section the maps produced under the Canadian Agri-Environmental Indicators 

programme (Clearwater et al., 2015) for wind and water erosion, SOC change and nutrient imbalance (residual 

soil N and P-source risk classes) are presented so that the linkage between drivers and threats to soil functions 

can be illustrated. The maps focus on agricultural impacts on soil functions. No comparable products exist for 

other soil threats. 

The major changes to drivers in Canada have been discussed in detail in Section 14.3 above. There are 

distinct differences across the main ecoregions of Canada that experience concentrated human impact. The 

main drivers can be summarized as: 

• Intensification of agriculture e.g. reduced pasture area, increased area of maize and soybean production, 

higher fertilizer inputs in the Mixed Wood Plains ecoregion in southern Ontario and Quebec and in the 

agricultural areas of eastern Canada e.g. in New Brunswick, Nova Scotia, and Prince Edward Island. 

• Significant reductions in the area of tillage summerfallow in the Temperate Plains and West-Central 

Semi-Arid ecoregions. 

• Widespread adoption of conservation tillage practices in most cropping systems in Canada. 

14.5.1 | Water and wind erosion 

In Canada, the Soil Erosion Risk Indicator was used as part of the Agri-Environmental Indicators programme 

to assess the risk of soil erosion from the combined effects of wind, water and tillage on cultivated agricultural 

lands (Figure 14.5). This indicator and its component indicators for wind, water and tillage erosion reflect the 

characteristics of the climate, soil and topography and correspond to changes in farming practices over the 

30-year period from 1981 to 2011. Wind and water erosion are the primary focus of this summary. For details on 

calculation of the indicators, see Li et al. (2008), McConkey, Li and Black, (2008), and Huang and Lobb (2013). 

The erosion indicator calculation estimates the rate of soil loss. These values are reported in five classes. 

Areas in the very low risk class are considered capable of sustaining long-term crop production and maintaining 

agri-environmental health under current conditions. The other four classes represent the degrees of risk of 

unsustainable conditions that call for soil conservation practices to support crop production over the long 

term and to reduce risk to soil quality. 

The risk of soil erosion on Canadian cropland has steadily declined between 1981 and 2011. The majority 

of this change occurred between 1991 and 2006. In 2011, 61 percent of cropland area was in the very low risk 

class overall, a considerable improvement over 1981 when only 29 percent was in this risk class. This decrease 

in water and wind erosion risk was most pronounced in the Temperate Prairie and West-Central Semi-Arid 

ecoregions in Alberta and Saskatchewan (Figure 14.5 and 14.6). Much of the improvement in erosion risk is 

from reductions due to the reduction in tillage summer fallow. A second driver is the increased adoption of 

direct seeding and conservation tillage, which is largely responsible for the decrease in tillage intensity and 

soil erosion. 


460 

in North America

Figure 14.5 Risk of water erosion in Canada 2011. Source: Clearwater et al., 2015. 


461 

in North America

Figure 14.6 Risk of wind erosion in Canada 2011. Source: Clearwater et al., 2015. 

Of the cropping systems across Canada, the risk of soil erosion by water is greatest under potato production 

in central and eastern Canada. In these areas there is intensive tillage and little opportunity to reduce the 

intensity through conservation tillage practices (Figure 14.5). The cropping system with the next greatest risk 

of erosion is the production of maize and soybeans under conventional tillage; however, there is a significant 

opportunity to reduce this erosion risk with conservation tillage. Of all soil landscapes across Canada, the 

risk of soil erosion by water is greatest in areas with maximum slopes of 10 percent or more, especially those 

located in central and eastern Canada where the risk of water erosion is inherently high due to climate 

(Figure 14.5). 


462 

in North America

Figure 14.7 Soil organic carbon change in Canada 201. Source: Clearwater et al., 2015. 


The soil organic carbon (SOC) change indicator used in the national Agri-Environmental Indicators 

programme assesses how organic C levels are changing over time in Canadian agricultural soils. The indicator 

is based on the method used for the Canadian National Inventory Report (Environment Canada, 2014). The 

indicator uses the Century model (NREL, 2007) to predict the rate of change of organic C content in Canada’s 

agricultural soils due to the effects of land management change since 1951. These include changes in tillage 

and summer fallow frequency, and change between annual crops and perennial hay or pasture. It includes 

land use changes such as clearing forests for agriculture or breaking native grass for cropland, but does not 

include the loss of C from the above-ground forest biomass. 

No changes in SOC were assumed if there were no indicated changes in land use or land management. The 

SOC change indicator does not consider soil erosion. 

The SOC change indicator results are presented (Figure 14.7) as the percentage of total cropland that falls 

into each of five SOC change classes expressed in kilograms per ha per year (kg ha -1 yr -1 ). Negative values 

represent a loss of SOC from the soil and positive values represent a gain of SOC. 


463 

in North America

For the Boreal Plains ecoregions from Ontario eastward, there was an overall loss of SOC from 1981 to 2011 

due to the reduction in the area of hayland and pasture and the corresponding increase in the area of annual 

crops (Figure 14.7). This shift in land use reflects a reduction in the demand for feed associated with the declining 

cattle populations in those provinces. The losses in Ontario and Quebec have been offset to a limited degree as 

a result of the adoption of conservation tillage. However, conservation tillage has not been implemented to 

the same extent in provinces in eastern Canada due to their cooler and wetter climatic conditions. 

The two agricultural ecoregions in the Prairie provinces have seen major increases in SOC over time due to 

reductions in tillage intensity and in summer fallow. These changes are responsible for the overall net gain in 

SOC in Canada (see Section 14.4.3 above). 


The assessment of nutrient imbalance in the Agri-Environmental Indicators programme focuses on N and 

P, and assesses both the N and P status of soils and the risk to water quality associated with the soil stores. 

The risk to water quality involves coupling hydrological and climate data with the land surface information for 

each region. This section focuses on the N and P status of soils. 

The residual soil nitrogen (RSN) indicator used in the National Agri-Environmental Indicators program 

provides an estimate of the amount of inorganic N that is left in the soil at the end of the growing season 

which may be susceptible to loss (Drury et al., 2007, 2010). The RSN indicator is estimated as the yearly 

difference between the total N input to agricultural soils and the output in harvested crops and gaseous losses 

including ammonia, nitrous oxide and dinitrogen. The major categories of N inputs into soil include fertilizer 

addition, manure application, biological nitrogen fixation by leguminous crops and free-living bacteria, and 

atmospheric wet and dry deposition. Nitrogen outputs include N removal in the harvested crop and gaseous 

N emissions via ammonia volatilization (NH 3 

), nitrification (N 2 

O) and denitrification (N 2 

O, N 2 

). 

The RSN on Canadian agricultural land has steadily increased from a low of 9.4 kg N ha -1 in 1981 to a maximum 

of 25.3 kg N ha -1 in 2001 (a year where many regions experienced drought conditions and were unable to use 

the applied N). The latest figure is 23.6 kg N ha -1 in 2011, the most recent census year. On a national basis, N 

inputs have almost doubled over the 30 years from 44.4 kg N ha -1 to 80.8 kg N ha -1 whereas N outputs have only 

increased by 63 percent from 35 kg N ha -1 in 1981 to 57.2 kg N ha -1 in 2011. 

The RSN map (Figure 14.8) for Canadian farmland in 2011 generally shows that there are high or very high 

residual N contents in farmland in many areas across Canada. Considerable change has occurred since 1981; 

for example, the Temperate Prairies in Manitoba were primarily in the very low and low risk classes in 1981 

whereas the great majority of the province is now in a very high risk class in 2011 (Figure 4.4). The Mixed Wood 

Plains in central and eastern Canada are also currently predominantly in the very high risk group, which again 

a considerable increase since 1981. 


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in North America

Figure 14.8 Residual soil N in Canada 2011. Source: Clearwater et al., 2015. 

Changes to management practices are required especially in the more humid regions in the Canada (notably 

Ontario, Quebec and eastern Canada) to reduce N losses from soils, to increase fertilizer use efficiency and to 

better synchronize N application with crop N demand. Further, the use of cover crops, especially in years with 

reduced yields, may help to reduce N losses from soil. 

The IROWC-P Indicator was developed to assess the status and trends over time for the risk of surface 

water contamination by P from Canadian agricultural land and is reported for agricultural watersheds (van 

Bochove et al., 2010). The initial stage in calculating IROWC-P involves the estimation of the annual amount 

of dissolved P that may potentially be released from agricultural soils (P source). P source is estimated as a 

function of cumulative P additions and removals (P-balance) over a 35-year period up to 2011 and the resulting 

degree of soil P saturation. 

There has generally been an increasing trend in the P-source levels in the surface of agricultural soils in 

Canada since 1976 as intensified agricultural practices have resulted in the application of P in excess of crop 

uptake (also called positive annual P balance) and have therefore increased soil P saturation. In 2011, very high 

concentrations of P (more than 4 mg of P per kg, or >4 mg P kg-1) at risk for release by storm events were located 

in regions where the agricultural production has been historically intensive and where soils have reached high 

P saturation values. High risk of water- contamination by P occurs around Abbotsford, British Columbia in the 

Marine West Coast Forest ecoregions, in the Temperate Prairies around Lethbridge in Alberta, and in portions 

of southern Saskatchewan and Manitoba (Figure 14.9). Intensive livestock operations near Abbotsford and 


465 

in North America

Figure 14.9 Indicator of risk of water contamination by phosphorus (IROWC-P) in Canada in 2011. Source: Clearwater et al., 2015. 

Lethbridge are major local sources of high P loadings. The Mixed Wood Shield ecoregion of central Ontario and 

the Mixed Wood Plains ecoregion in Quebec, New Brunswick and Prince Edward Island are also dominated by 

very high or high P source risk. 

Implications of soil threats for soil functions 

Clearly the national assessment of the threats to soil functions shows a distinct separation between the 

agricultural systems of the Prairie provinces and those of Ontario, Quebec, and Atlantic Canada. Both the 

state and the trend of soil change in the Prairie provinces are generally positive, especially in soil erosion 

(notably wind erosion) and carbon change. The greatest risk in this region lies in the high residual nutrient 

levels in areas like Manitoba and in the possible contribution of these high residual levels to eutrophication in 

lakes in this region (Schindler, Hecky and McCullough, 2012). 

The level of the threats in the Mixed Wood Plains of central and eastern Canada is very different. There 

are generally high or very high levels of threat for soil organic carbon change, erosion by water, and nutrient 

imbalance. The well-integrated drainage system and higher precipitation levels than in the Prairie provinces 

lead to significant sediment and nutrient delivery to waterways. There is thus a risk of serious impact of 

agricultural land management on water quality (Clearwater et al., 2015). The combined effects of soil organic 

carbon loss and water erosion presumably also reduce services and products delivered by the soil, but these 

have been poorly documented in this region. 


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14.6 | Conclusions and recommendations 

Overall there has been significant progress made in reducing threats to soil functions in North America. 

Threats from acidification and contamination have been reduced due to the imposition of a stronger 

regulatory framework. The greatest change in cropping practices - reduced tillage - has largely come about 

through adoption by individual producers supported by government and private sector extension agents. 

However, major areas of concern remain. Erosion rates are still above what are believed to be tolerable levels 

in the Temperate Prairies ecoregion of the United States and throughout the Mixed Wood Plains ecoregion of 

Canada and the United States. Transport of soil-derived N and P to waterways is a major problem, and excess 

application of N continues throughout much of the cropland in the United States and in central and eastern 

Canada. 

Although a wide variety of best management practices for optimum nutrient application and erosion 

control have been developed and promoted, the problems of erosion and nutrient imbalance persist. 

Salinization, contamination, and acidification affect smaller areas in North America, and in the case of the 

latter two threats the current regulatory framework limits the expansion of the affected area. Waterlogging 

is little studied in North America, and we recommend that future reports include assessments of loss of 

wetlands as an important metric for sustainable management in this and other regions. 

The loss of agricultural land to soil sealing is not perceived as a major issue in North America. However, the 

paucity of data on this threat needs to be addressed for a more informed assessment to be made. 

Changes in carbon stocks in North America have been extensively modeled as part of national reporting 

programs on greenhouse gas emissions, but only in a few landscapes are the models adequately supported 

by field observations of SOC change. The greatest uncertainty surrounding SOC change lies in the response of 

carbon in permafrost soils to climate change in northern Canada and Alaska and improved monitoring of this 

response is essential. 

Like the SOC models, the agri-environmental indicators approach used in the Canadian case study allows 

estimation of the change in threats to soil function over time. However, several criticisms can be made of 

the approach. First, there is a lack of ongoing monitoring of important soil physical, chemical and biological 

properties at relevant scales through time and hence it is difficult to assess the performance of the models 

that underlie the indicators. Second, the interaction between indicators (for example, between erosion and 

carbon) is not considered, and this can bias the assessment of soil change. Third, it is difficult to assess the 

variability associated with modelled results and therefore to evaluate overall the confidence that can be placed 

in the results. Overall there is a need to revise the models in light of new scientific advances and to develop and 

refine scientifically credible programs to assess model performance. 

The greatest uncertainty overall in our knowledge about the threats to soil functions lies in our limited 

understanding of the changes in soil biodiversity in the past and present and the implications of these changes 

for sustainable soil management. 

Based on the above finding, a provisional assessment is made of the status and trend of the 10 soil threats in 

order of importance for the region. At the same time an indication is given of the reliability of these estimates 

(Table 14.1) 


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in North America


function 


Nutrient 

imbalance 

Organic 

carbon change 

Loss of soil 

biodiversity 

Summary 

Reduced tillage and 

improved residue 

management have lowered 

erosion rates in regions 

such as the Great Plains in 

Canada but water erosion 

rates continue to be too 

high in the northern 

Mid-West of the U.S. and 

agricultural areas of central 

and Atlantic Canada. 

Excess application 

of fertilizers in many 

regions causes significant 

degradation of surface 

water quality and increased 

emissions of nitrous 

oxide to the atmosphere. 

Contamination of surface 

water is strongly linked 

to high erosion rates, and 

occurs in the same regions 

(northern midwest U.S., 

Mississippi River Basin, 

and agricultural regions of 

central Canada). 

The majority of cropland 

in the U.S. and Canada 

has shown improvements 

in SOC stores due to the 

wide-spread adoption of 

conservation agriculture 

(i.e., reduced tillage 

and improved reside 

management). There is a 

lack of field validation sites 

to support the nationallevel 

modelling results. 

Loss of SOC from northern 

and Arctic soils due to 

climate change is a major 

concern. 

The extent of loss of soil 

biodiversity due to human 

impact is largely unknown 

in North America. The 

effects of increasing 

agricultural chemical use, 

especially pesticides, use 

on biodiversity is a major 

public concern. 

Known level of carbon loss 

suggests similar loss in 

biodiversity. 



Confidence 

In 

In trend 

condition 


468 

in North America

Compaction 

Sealing 

and land take 

Salinization 


Contamination 


acidification 

Waterlogging 

Compaction continues 

to be a low-level issue, 

especially in regions with 

texture-contrast (Luvisol, 

Alfisol, Ultsol) soils. The 

regional-scale impact 

of compaction on plant 

growth is largely unknown. 

Substantial expansion of 

housing and infrastructure 

in areas of high quality 

farmland continues in 

both countries but is 

(incorrectly) not perceived 

as a concern. Neither 

country has reliable data 

on sealing and land take. 

Salinization is believed to 

be increasing in parts of 

the northern Great Plains 

in the U.S.A. but the risk of 

salinization is decreasing in 

western Canada. 

Although many legacy 

contamination sites exist, 

improved regulatory 

systems in both 

countries has limited the 

creation of new areas of 

contamination. 

Large-scale land 

disturbance due to 

resource extraction 

activities continues to be a 

significant issue. 

Trans-national 

environmental legislation 

has significantly reduced 

soil acidification in forested 

areas of eastern and 

central North America. 

Localized areas of 

acidification in agricultural 

land managed through 

lime application. 

Waterlogging is not 

believed to be a significant 

threat in North America. 

Localized flooding has 

occurred due to a wider 

amplitude of precipitation 

events in the past decade. 

Loss of wetlands is a more 

significant threat 

in North America. 

Table 14.1 Summary of soil threats status, trends and uncertainties in North America 


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in North America

Acknowledgements 

Jenny Sutherland of NRCS in Lincoln, Nebraska very capably edited the material on the soil threats. Tim 

Martin of Agriculture and Agri-Food Canada provided access to the updated Canadian agri-environmental 

indicators material used in this chapter. The chapter was reviewed by D. Burton, F. Walley, C. Rice, E. Gregorich, 

C. van Kessel, T. Moore, and F. Larney and their suggestions for revisions are gratefully acknowledged. 

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15 | Regional Assessment of Soil 

Change in the Southwest Pacific 

Regional Coordinator/Lead Author: N.J. McKenzie (ITPS/Australia) 

Contributors: J.A. Baldock (Australia), M.R. Balks (New Zealand), M. Camps Arbestain (ITPS/New Zealand), 

L.M. Condron (New Zealand), M. Elder-Ratutokarua (Secretariat Pacific Community/Fiji), M.J. Grundy 

(Australia), A.E. Hewitt (New Zealand), F.M. Kelliher (New Zealand), J.F. Leys (Australia), N.J. McKenzie (ITPS/ 

Australia), R.W. McDowell (New Zealand), R.J. Morrison (Fiji/Australia), N.R. Schoknecht (Australia). 


476 

in the Southwest Pacific


The Southwest Pacific region includes the 22 island nations of the Pacific1, New Zealand and Australia 

(Figure 15.1). The landscapes of the region are very diverse ranging from a large continental land mass through 

to tens of thousands of small islands across the enormous expanse of the southwest Pacific Ocean. There 

are extensive ancient flat lands through to some of the youngest and most tectonically active landscapes on 

the planet. Temperature and rainfall ranges are large because of the breadth of latitudes and elevations. As a 

consequence, the soils of the region are also diverse. The strongly weathered soils in humid tropical areas and 

the vast expanses of old soils across the Australian continent are particularly susceptible to disturbance and 

this is where some of the more intractable problems of soil management occur today. 

15.2 | The major land types in the region 

The major land types in the region owe their origin to the relative movement of the Earth’s lithospheric 

plates, and in particular to the interaction between the Australian and Pacific Plates. The breakup of the 

supercontinent of Gondwana included the separation, around 96 million years ago, of the Australian Plate 

from the Antarctic Plate. The Australian Plate includes the present day island continent of Australia, Papua 

New Guinea, a small part of the South Island of New Zealand, and some islands including New Caledonia 

and Norfolk Island. The Australian Plate is moving northwards at approximately 70 mm yr -1 , colliding with the 

Eurasian Plate. This has created the mountains of Indonesia and Papua New Guinea, the highest peaks being 

Puncak Jaya (4 848 m) and Mt Wilhelm (4 509 m) respectively. The Pacific Plate is moving westwards towards 

the Australian Plate. Movement along the transform boundary of these plates has created the Southern Alps 

of New Zealand (Aoraki/Mt Cook 3 724 m) and related geological activity has resulted in substantial volcanic 

activity around the aptly named Pacific Rim of Fire. New Zealand straddles the boundary of the Australian and 

Pacific Plate. Most of the island nations of the Pacific are volcanic in origin and exist today as either hilly or 

mountainous features formed by the volcanoes themselves, or as atoll islands. 

1 These countries are American Samoa, Cook Islands, Federated States of Micronesia, Fiji, French Polynesia, Guam, Kiribati, Marshall Islands, Nauru, New 

Caledonia, Niue, Northern Mariana Islands, Palau, Papua New Guinea, Pitcairn Islands, Samoa, Solomon Islands, Tokelau, Tonga, Tuvalu, Vanuatu, and Wallis 

and Futuna. 


477 


Figure 15.1 Nations in the Southwest Pacific region and the extent of Melanesia, Micronesia and Polynesian cultures. Figure based on 

base map imagery: exclusive economic zone boundaries (EEZ)v 8 2014, Natural Earth 11 3.2.0 

Ancient landscapes 

After separating from Antarctica, the Australasian landmass moved northwards into warmer latitudes at 

the same time that the globe was cooling during the Pleistocene. The development of the circumpolar Southern 

Ocean further moderated the climate of the Australasian landmass. The resulting relative stability meant 

that biological evolution and soil development occurred on similar timescales and without major phases of 

interruption. This is in contrast to much of the Northern Hemisphere where repeated glaciations renewed 

landscapes and ensured that large areas have relatively young soils. This long history of soil development has 

many implications today for land management in the region. 

Low-relief landscapes of Australasia 

The western two-thirds of the Australian continent are dominated by ancient landscapes and strongly 

weathered soils. Some of these soils bear the imprint of previous climates with some unexpected patterns of 

soil distribution – very acid leached soils (normally associated with humid regions) now occur in deserts, and 

deeply weathered soil profiles (tens of metres deep) occur in Mediterranean climates with limited leaching. 

Vast areas of sandy soils have formed from the predominantly acid-igneous and sedimentary parent materials. 

Nutrient status is very poor and micronutrient deficiencies are common. In the south, substantial quantities 

of salt have accumulated in the low relief landscapes. As a consequence, sodic and saline soils are widespread 

and human-induced salinity is a major land management problem. 

Uplifted and eroded continental margins 

The soils and landscapes of the eastern third of Australia are dominated by the influence of the Great 

Escarpment (Ollier, 1982) – a landscape feature that extends for 3 000 km from Northern Queensland to 

Victoria. The Great Escarpment was formed by uplift associated with the passive continental margin to the 

east. Inland of the Escarpment, the soils and landscapes tend to be older but more clay-rich than in the west 


478 


of the country. The undulating and low relief landscapes also have saline and sodic soils. Large sedimentary 

basins (e.g. Murray–Darling Basin) have moderately fertile soils although the current climate is highly variable. 

The elevated tablelands adjacent to the Great Escarpment have significant areas of basalt and their 

associated alluvial landscapes have some of Australia’s best soils for agriculture. Likewise, the dissected 

landscapes to the east and south of the Great Escarpment where most Australians live are much younger. The 

coastal river systems have fertile alluvial landscapes. 

The remnants in New Caledonia, Papua New Guinea and New Zealand 

The areas of the Australian Plate in New Caledonia, Papua New Guinea and New Zealand share ecosystems 

with a common evolutionary origin (e.g. Antarctic flora including Araucaria and Nothofagus). They typically 

have high levels of biodiversity although the above-ground systems are much better documented that those 

in the soil. These remnant areas are on the more tectonically active fringes of the Australian Plate and they 

have greater relief as a result. 

Young active landscapes 

New Zealand 

The tectonically active landscapes of New Zealand can be broadly divided into the axis of high mountain 

ranges, the basin and range provinces on either side of these ranges, the Taupo Volcanic Zone on the North 

Island, and the lowlands and sedimentary basins. Only 15 percent of the land area is flat. The climate of the 

country ranges from sub-tropical in the north to sub-Antarctic in the far south. Sixty percent of New Zealand 

is >300 m above sea-level and there are about 3 000 mid-latitude mountain glaciers. 

A climate sequence of Luvisols, dystric Cambisols and Podzols cover 69 percent of the country and are 

derived from sedimentary rocks (greywacke, sandstone, siltstone and mudstone as colluvium, alluvium and 

loess). Significant soils in the remaining areas include vitric Andosols derived from rhyolytic tephra (North 

Island) and silandic Andosols, Nitisols and Ferralsols derived from andesitic tephra and basalt (mainly on the 

North Island). 

Papua New Guinea 

Papua New Guinea has five major landscape regions (Löffler, 1977, 1979). The Southern Plains and Lowlands, 

up to 400 km wide, are in the west of the country. Most of this region is less than 30 m above sea level and 

includes the extensive alluvial plains of the Fly River. There are two mountainous regions that occupy the 

majority of the country. The Central Ranges run the length of the mainland and have very high relief with many 

peaks between 3 000 m and 4 000 m. The Northern Ranges run parallel and descend to a discontinuous 

and narrow coastal plain. Between these ranges are the plains, lowlands and wetlands of the Inter-montane 

Trough, many of which are associated with the Sepik River. The Islands Region is diverse and includes active 

volcanoes (e.g. New Britain) particularly along the Northern Bismarck Island Arc. Fringing coral reefs and 

raised coral limestone landforms are common in this tectonically active area. 

Island landscapes and atolls 

As noted earlier, most of the smaller islands in Melanesia, Micronesia and Polynesia are volcanic in origin. 

Those with hilly or mountainous features were formed by the volcanoes themselves and they may have 

fringing coral reefs, elevated coral platforms, or both. The low lying atoll islands have been formed by corals 

growing on extinct seamounts or volcanoes that have eroded or subsided. Many of the atoll islands are only a 

few meters above sea level and are therefore vulnerable to sea-level rise. 


479 


15.3 | Climate 

The climate of the region is strongly influenced by circulation patterns and processes in the Pacific Ocean 

(the Southern Oscillation) that bring La Niña conditions associated with floods and cyclones and El Niño 

conditions that are associated with droughts. A similar circulation pattern in the Indian Ocean (the Indian 

Ocean Dipole) influences drought across southern Australia. The Southern Annular Mode of the Southern 

Ocean affects weather and climate in New Zealand and southern Australia. These large-scale circulation 

processes interact with the landscape (e.g. through orographic processes and in accord with the scale of 

the land mass) to exert a strong control on land use and management across the region. Some of the most 

significant features are as follows. 

• Australia has a very highyear-to-year rainfall variability, and major droughts and wet periods occur on 

a decadal scale. Resilient systems of land and water resource management are essential to deal with 

this level of climate variability. 

• Landscapes in the tropics and sub-tropics experience cyclones and very high intensities of rainfall 

especially in coastal areas. Maintenance of surface cover is essential to avoid extremely high rates of 

soil erosion. 

• The wet-dry tropics of northern Australia and some Pacific Islands receive most of their rainfall in three 

consecutive months and the remainder of the year is severely water-limited. This restricts options for 

land use unless some form of irrigation is possible. The harsh climate, in conjunction with the ancient 

and strongly weathered soils, largely accounts for the dramatic difference in land use and population 

density of Northern Australia when compared to nearby Indonesia and Papua New Guinea. 

In some parts of the region, there is a sensitive interplay between rainfall, evaporation and the capacity 

of soil to store water. Many soils in southern Australia have a limited capacity to store water and this makes 

them especially vulnerable to small changes in the distribution and amount of rainfall. As a consequence, 

relatively small changes in rainfall and temperature caused by climate change are having a significant impact 

on farming systems and water resources (Reisinger et al., 2014). Likewise, sea-level rise caused by global 

warming (Nurse et al., 2014) creates an immediate and serious threat to thousands of low lying atoll islands 

in the region. 

15.4 | Land use 

15.4.1 | Historical context 

The history of human settlement has occurred in several distinct waves over a very long period. In every 

case the impact on soils has been substantial and in some cases catastrophic. The earliest records indicate 

that human arrival occurred in Papua New Guinea and Australia at least 45 000 years BP. At this time, sea 

levels were much lower and a land bridge connected the two countries forming the single continent of Sahul. 

This continent was widely colonized by 35 000 years BP (O’Connell and Allen 2004). 

It is difficult to assess the impact on soils caused by the initial colonization of the region, particularly by 

the Australian Aborigines, because it coincides with a period of rapid climate change towards the end of the 

Pleistocene. In Australia, major changes in fire, vegetation, and wildlife occurred (Roberts, Jones and Smith, 

1990; Roberts et al., 1994; Bowler et al., 2003; Turney et al., 2001). The cumulative effect on soils caused 

by humans and other environmental drivers during this time may rival the later direct impact caused by 

Europeans, although the latter has been concentrated into a very short period. 


480 


The broad area of Near Oceania (including present-day Papua New Guinea, Solomon Islands, Vanuatu, 

New Caledonia and Fiji) was occupied by peoples that were to remain relatively isolated for more than 25 000 

years giving rise to the remarkably diverse cultures and languages of Melanesia – for example, more than 800 

languages are still spoken today in Papua New Guinea. It is likely that agriculture was invented in the New 

Guinea highlands at about the same time (10 000 years BP) as it appeared in other parts of the world, and 

that agricultural development and plant domestication in New Guinea was independent of what happened 

elsewhere in the world until about 3 500 years BP (Denham et al., 2003). 

The Polynesian and Micronesian peoples arrived in relatively recent times and their ancestors were 

primarily from East Asia. The development of sailing approximately 3 000 years BP enabled their colonization 

of the islands in Remote Oceania as far east as Tonga and Samoa where Polynesian culture then developed 

(Friedlaender et al., 2008). 

When the Polynesian Maoris first arrived in New Zealand c. 1200 CE, most of the land was covered by dense 

forest which they began to clear by burning to facilitate settlement and hunting (McGlone, 1989). Most of the 

forest clearance occurred between 1350 and 1550, with impacts on soils and landscapes that were greatest in 

the drier east coast regions of both main islands. By 1840 just prior to large scale European settlement, the 

forest cover had been reduced from 85 percent (23 million ha) to 53 percent (14.3 million ha) (Condron and Di, 

2002). 

15.4.2 | Nineteenth and twentieth centuries 

European colonization was widespread across the region during the nineteenth and twentieth century. The 

impact on soils in many districts, particularly in Australia and New Zealand, was profound and in some areas 

initially catastrophic. In Australia, the severity of soil degradation, particularly in the 100 years after 1850, was 

extreme, culminating in the dust bowl years of the 1930s and 1940s. In New Zealand, the clearing of steep hill 

country led to widespread erosion and sedimentation in river systems. 

Other countries in the region (e.g. Papua New Guinea, Fiji, French Polynesia) went through distinct phases 

of land use in this period. Subsistence shifting-agriculture was disrupted by European commercial farming 

(e.g. pastoralism, coconut plantations, cotton and coffee) which often involved widespread clearing of the 

tropical lowland forests. Sugar cane industries were established in some countries and the logging of the 

indigenous timber resource was widespread. Urban areas expanded and in some countries agriculture spread 

onto more marginal lands with intensification of land use being widespread. 

The legacy of the early and destructive phases of land use is still evident throughout the region. Many 

landscape processes take decades to stabilize after a change in land use and as a consequence, the full extent 

of soil change caused by prior land use across the region has yet to be fully expressed. The most significant 

and potentially irreversible causes of soil change across the region are discussed in the following paragraphs. 

Clearing: In Australia, the removal of deep-rooted native vegetation and its replacement with annual crops 

and perennial pastures has in most cases led to a net loss of organic carbon, nutrients and less efficient use by 

plants of the available rainfall. This can lead to rising groundwater levels and, in some cases, to dryland salinity. 

The time between initial disturbance and the longer term equilibrium for different soil properties can range 

from a few decades (e.g. organic carbon in light-textured soils (Sanderman, Farquharson and Baldock, 2010) 

through to many decades or centuries (e.g. dryland salinity in regional-scale groundwater systems (NLWRA, 

2001b)). Conversely, in New Zealand, carbon and nitrogen stocks increase (in particular in hill country with 

low stocking rates) while C/N ratios decline when converting from native vegetation to pasture (Schipper and 

Sparling, 2011; Sparling et al., 2014). 


481 


Mining: Major disruptions to landscapes were caused by mining, beginning with the gold-rush era (from 

the 1850s onwards) in Australia and New Zealand. Alluvial landscapes in particular were dug up, turned over, 

sluiced and degraded. Enormous quantities of sediments were generated and soils and river systems were 

damaged irreparably. The environmental controls around mining operations in these two countries have 

improved in more recent decades but this has not been the case elsewhere in the region. 

Mining operations on various Pacific islands have had major environmental and social impacts. Examples 

include the Ok Tedi (see below) and Porgera mines in Papua New Guinea, extensive strip mining for phosphorus 

in the small country of Nauru where >90 percent of the original soils have been removed (Morrison and 

Manner, 2005), and the ill-fated Bougainville Copper mine. 

During the last decade, the area of land used for mining and extraction of energy resources (e.g. coal and 

coal seam gas) has expanded significantly in eastern Australia. This has generated conflicts and difficult policy 

decisions on land use because some of the highest potential areas for coal and coal seam gas coincide with 

high-quality agricultural land (Chen and Randall, 2013). 

Excessive cultivation and compaction: This was widespread during the first half of the 20th century and 

it is still a problem in some parts of the region. The economic cost of compaction caused primarily by heavy 

machines and animals is likely to be large but it remains unquantified. 

Grazing: Excessive grazing damages soil quickly because removal of cover leads directly to erosion. It 

can also permanently remove nutrients unless replenished with fertilizer. Large areas across the Australian 

rangelands have been affected (Bastin, 2008). In New Zealand, conversion of forested hill country to pastures 

has had a large impact on soil erosion. 

Rabbits and other feral animals: Wild grey rabbits were introduced to Australia in 1860 and within a decade 

they reached plague proportions. Their main effect was to exacerbate over-grazing caused by stock, leading 

to bare ground and erosion by water and wind. Rabbit plagues continued throughout the first half of the 20th 

century and were particularly devastating during the 1940s. Numbers were reduced greatly when in 1950 the 

myxomatosis virus was introduced. The pest gradually increased again in the following decades but was again 

reduced when the calici virus was released in the 1990s. Rabbits and other feral animals (e.g. goats and pigs) 

degrade soils in other parts of the region. They are a significant problem in New Zealand and on sub-Antarctic 

islands such as Macquarie Island which has been severely eroded in recent decades because of rabbits. 

15.4.3 | Contemporary land-use dynamics 

At the end of the twentieth century, diverse systems of land use operated throughout the Southwest Pacific 

ranging from traditional systems (e.g. subsistence shifting-agriculture in Papua New Guinea and some of the 

Pacific Islands) through to technologically advanced systems (e.g. highly mechanized agriculture in Australia 

and New Zealand). While the general patterns of land use have been relatively stable for several decades, 

significant changes are continuing. Those with the greatest impact on soil resources are summarised in Table 

15.1. There are large differences in the degree of economic development in the region (Table 15.2) as the region 

contains some of the world’s poorest nations and some of the richest. There are also large differences within 

nations. Twenty percent of Australia is owned and managed by economically disadvantaged Indigenous 

Australians (SOE, 2011). Approximately 5.6 percent of New Zealand is Maori land, a major contributor to 

New Zealand’s economy. About 26 percent of Maori businesses are in the primary sector, principally farming, 

forestry and horticulture. In some countries (e.g. Solomon Islands and Vanuatu) increasing rural populations 

have led to significantly greater use of productive land in the subsistence shifting-cultivation systems with 

much reduced fallow periods and subsequent declines in soil quality. 

Primary industries play a major role in most countries. The development of mineral and energy resources 

has been a key driver of economic growth, particularly in Australia and Papua New Guinea. New Zealand and 

Australia are major exporters of agricultural products. 


482 


Country and land-use driver 

Implication for soil resources 

Australia 

Commodity prices 

Cessation of land clearing 

Expansion of mining and energy 

industries 

Agricultural land management 

Population growth and urban 

expansion 

Drying of southwest Western 

Australia 

Land degradation 

Disruption to, or intensification of, current systems of land use and 

management 

Carbon stocks in areas under native vegetation are protected. 

Intensification of land use occurs elsewhere 

Loss of productive soils used for agriculture 

Intensification of land use may increase some threats (e.g. 

acidification, nutrient imbalance) but improved systems of soil 

management are reducing others (e.g. compaction, carbon loss, 

contamination) 

Sealing and capping of land 

Risks to farming systems, forestry and other plantations 

Decreased viability of current systems of land use (e.g. southern 

rangelands of Australia) 

New Zealand 


Intensification of agriculture 

Expansion of forestry 

Urban expansion 

Disruption to, or intensification of, current systems of land use and 

management 

Intensification of land use may increase some threats (e.g. nutrient 

imbalance) but improved systems of soil management are reducing 

others (e.g. compaction, carbon loss, contamination) and other 

off-site impacts are likely (e.g. eutrophication of waterways, 

increased emissions of GHG) 

Reduced risk of erosion and potential changes in carbon stocks 


Papua New Guinea 

Rapid population growth 

Unsustainable logging 

Mining 

Reduced rotation length in traditional subsistence systems leading 

to fertility decline 

Erosion, fertility decline, reduced carbon stocks, loss of biodiversity 

Immediate site impacts and risks of off-site contamination 

Fiji 


Forest management 


Disruption to established systems of land use and management 

(particularly for the sugar industry) 

Risks of erosion, fertility decline, reduced carbon stocks, loss of 

biodiversity 

Sealing and capping of versatile land 


483 


Solomon Islands 

Unsustainable logging 

Erosion, fertility decline, reduced carbon stocks, loss of biodiversity 

Low Lying Atoll Islands 

Local production of fresh food 

Sea-level rise 


Improved management of limited soil resources 

Salinization, loss of soil resources 


Table 15.1 Summary of current primary drivers of land-use and the associated implications for soil resources in the Southwest Pacific 

region. 

Country 

Population 2015 

Thousands 

Population 2050 

Thousands 

GDP per capita USD 

(2009-2013) 

Australia 23 923 33 735 67 468 

New Zealand 4 596 5 778 41 556 

Melanesia 

Fiji 893 918 4 572 

New Caledonia 263 364 - 

Papua New Guinea 7 632 13 092 2 088 

Solomon Islands 584 1010 1 954 

Vanuatu 264 473 3 303 

Total Melanesia 9 636 15 858 

Polynesia 

American Samoa 56 62 - 

Cook Islands 21 24 - 

French Polynesia 283 337 - 

Niue 1 1 - 

Samoa 193 242 3 647 

Tokelau 1 1 - 

Tonga 106 140 4 427 

Tuvalu 10 12 3 861 

Wallis and Futuna Islands 13 13 - 

Total Polynesia 684 832 

Micronesia 

Federate States of Micronesia 104 130 3 235 

Kiribati 106 156 1 651 

Marshall Islands 53 67 3 325 

Nauru 10 11 - 

Northern Mariana Islands 55 52 - 

Palau 21 28 11 810 

Total Micronesia 519 671 

Table 15.2 Current population, project population (UNDESA, 2013) and Gross Domestic Product per capita (World Bank, 2014) for 

countries of the region. 


484 


15.5 | Threats to soils in the region 

15.5.1 | Erosion by wind and water 

The rates of soil erosion occurring today in Australia and New Zealand are significantly less than in previous 

decades. The situation in the rest of the region is less clear. Very fast rates of erosion are occurring in countries 

with uncontrolled land clearing and logging (e.g. Papua New Guinea and the Solomon Islands). Unsustainable 

rates of erosion are also likely to be occurring in marginal and hilly lands used for agriculture in some Pacific 

countries, for example Fiji (Liedtke, 1989). 

Soil erosion by wind is a significant problem in Australia and an account of trends and current status is 

presented below. It is not common elsewhere in the region because of the humid climate, although drier areas 

in New Zealand (primarily on the South Island) are prone to wind erosion (Eyles, 1983). 

Australia 

Current rates of soil erosion by water in Australia are much less than the peak periods just after land clearing. 

In many parts of the country, widespread gully erosion occurred during this time and the hydrological regime 

of many river systems was changed. In southern Australia, gully and river bank erosion are the dominant 

sources of sediment supplied to streams. Gully erosion in southern Australia has now been largely stabilised, 

but gullies are still actively forming in northern Queensland and in some agricultural regions of Western 

Australia (NLWRA, 2001a). 

Despite the apparent stabilization, current rates of soil erosion by water across much of Australia now 

exceed soil formation rates by a factor of at least several hundred and, in some areas, several thousand. As 

a result, the expected half-life of soils (the time for half the soil to be eroded) in some upland areas used for 

agriculture ranges from less than a century to several hundred years. The latest assessment concluded that 

soil erosion by water in Australia is still at unsustainable rates, but there are large uncertainties about the time 

until soil loss will have a critical impact on agricultural productivity (SOE, 2011; Bui et al., 2010; Bui, Hancock 

and Wilkinson, 2011). Environmental impacts of excessive sedimentation and nutrient delivery on inland 

waters, estuaries and coasts are already occurring. 

It is estimated that up to 10 million ha of land have less than 500 years until the soil’s A-horizon (effectively 

the more fertile topsoil) will be lost to erosion. Most of this land is in humid subtropical Queensland. Integrated 

studies of soil formation and erosion using a variety of techniques will be needed to better understand 

the extent, severity and significance of the problem. However, it is clear that a concerted program of soil 

conservation is essential to control this chronic form of land degradation across large areas of Australia. The 

problem is arguably having its greatest environmental impact on the World Heritage listed Great Barrier Reef. 

The latest scientific consensus is that the decline of marine water quality associated with terrestrial runoff 

from the adjacent catchments is a major cause of the current poor state of many of the key marine ecosystems 

of the Great Barrier Reef. The main source of excess nutrients, fine sediments and pesticides from Great Barrier 

Reef catchments is diffuse source pollution from agriculture (Brodie et al., 2013). 


485 


Figure 15.2 Change in the percentage area of all land prepared for crops and pastures under different tillage practices in Australia, 

1996-2010 Source: SOE, 2011. 

Land management practices have improved significantly during the past few decades, due to better grazing 

practices, adoption of conservation tillage, enforcement of forestry codes and soil conservation measures in 

engineering (e.g. relating to road construction and urban development). 

Ground-based monitoring of management practices and land cover along with data on land-management 

practices (SOE, 2011) reveal a pattern of: 

• more careful grazing and maintenance of effective land cover at critical times of the year 

• improved adoption of conservation practices, especially across the cropping lands of southern Australia 

• an associated large decline in the amount of tillage in farming systems (Figure 15.2) 

New Zealand 

Soil loss by erosion is a major problem in many areas of New Zealand due to a combination of factors 

including soil type, topography and climate, as well as the type and associated intensity of land use (especially 

pastoral agriculture). 

Eyles (1983) provided the first systematic inventory of soil erosion in New Zealand and made the following 

observations and estimates. Surface erosion occurs mainly on the South Island. In 1983 sheet erosion 

affected 10 million ha of the country, while wind erosion affected 3 million ha (the total area of New Zealand 

is approximately 27 million ha). Mass movement occurred mainly on the North Island and slip erosion was 

estimated to affect 7 million ha. Fluvial erosion also occurred mainly in the North Island with rill and gully 

erosion estimated to affect 2 million ha. The study concluded that a total of 9 million ha of farmed land in New 

Zealand was at risk of significant erosion, although the level of risk was variable. The large scale afforestation 

of hill country and steep-land pasture since 1983 (see case study below) will have substantially reduced rates of 

soil erosion (especially sheet and slip erosion). However, forest harvesting results in significant soil disturbance 

and this will increase the risk of erosion when the first rotation forest is harvested and the second rotation 

established. 

Dymond (2010) analysed erosion rates in order to refine the national carbon budget because soil erosion 

in New Zealand results in the large export of sediment and particulate organic carbon (POC) directly to the 

sea. The North Island of New Zealand was estimated to export 1.9 (−0.5/+1.0) million tonnes of POC per year 

to the sea, and to sequester 1.25 (−0.3 /+0.6) million tonnes of carbon per year from the atmosphere through 


486 


egenerating soils. The South Island was estimated to export 2.9 (−0.7/+1.5) million tonnes of POC per year, and 

to sequester approximately the same amount. Dymond (2010) assumed that exported carbon is buried at sea 

with an efficiency of 80 percent. This gives New Zealand a net carbon sink of 3.1 (−2.0/+2.5) million tonnes per 

year (equivalent to about 45 percent of New Zealand’s fossil fuel carbon emissions in 1990). There is essentially 

a ’conveyor-belt’ transfer of carbon from the atmosphere to soils regenerating from erosion and to the sea 

floor where carbon is permanently buried. The large magnitude of the net sink is primarily due to tectonically 

driven uplift and erosion combined with high biological productivity. The degree to which other elevated and 

humid islands in the Pacific can operate as a net sink has not been determined. 

Pacific Islands 

The high rainfall and steep lands in many of the larger Pacific Island countries make them vulnerable to soil 

erosion. Understanding the significance of this erosion requires information on both rates of soil formation 

and the baseline rates of sediment movement in different landscapes. The latter can be much larger than in 

temperate regions and relatively fast rates of erosion around 50 tonnes ha -1 yr -1 have been recorded in heavily 

forested landscapes in Fiji (Glatthaar, 1988; Liedtke, 1989) and Samoa (Terry, Garimella and Kostaschuk, 2002; 

Terry, Kostaschuk and Garimella, 2006). Tropical cyclones, intense rainstorms, steep slopes and landslides are 

key factors. 

The desire for economic development has led to increasing areas of good quality soil being used for cash 

crops such as coffee and cocoa and for cattle production. These factors in conjunction with population 

pressures are forcing small-scale agriculture onto steeper less suitable lands. Logging, whether managed or 

illegal, increases the pressure even further. Condron and Di (2002) indicate that rates of soil loss by erosion in 

these areas are typically between 10 to 90 tonnes ha -1 yr -1 (see also Asquith, Kooge and Morrison, 1994; Liedtke, 

1989). Accelerated soil erosion has also been associated with plantations, particularly on marginal sloping 

land. 

Although most atolls are away from the cyclone belt in the Southwest Pacific, a few have been severely 

eroded by large storms. The impacts are devastating as much of the limited soil resource is washed away, and 

also what is left becomes highly salinized and of limited productive capacity. This was particularly noted on 

Funafuti, Tuvalu following cyclone Bebe in 1972 (Maragos, Baines and Beveridge, 1973). 

Sub-Antarctic Islands 

Several sub-Antarctic islands in the Southwest Pacific region continue to have high rates of soil erosion. 

Macquarie Island, for example, is an isolated island in the Southern Ocean that provides a critical habitat for 

migratory species. Macquarie Island has been severely eroded and approximately AUD $25 million has been 

spent in recent years on habitat restoration programmes, mainly for the eradication of rabbits. 


Most of the region experienced large losses of soil carbon when land was first cleared for agriculture. 

The major phases of clearing in each country were noted above. In some countries clearing associated with 

uncontrolled logging continues today, for example in Papua New Guinea and the Solomon Islands (Moorehead, 

2011). The initial disruption to soil caused by clearing usually results in a significant loss of nutrients. Organic 

matter is oxidised and the removal of surface cover (litter and protective vegetation) makes the soil more 

prone to erosion. Stores and cycles of nutrients adjust under the new land use, but in most cases the net 

loss of nutrients and leakage are greater than under natural conditions. Most studies indicate that across 

the region soil carbon typically reduces to 20–70 percent of the pre-clearing amount (e.g. Sanderman and 

Baldock, 2010). Opportunities for restoring some of this very large stock of carbon have been a focus of soil 

research in Australia and New Zealand during the last decade. 


487 


Australia 

In their review of replicated Australian field trials with time-series data, Sanderman and Baldock (2010) 

concluded that, although the implementation of more conservative land-management practices will lead to 

a relative gain in soil carbon, absolute soil carbon stocks may still be on a trajectory of slow decline. There are 

also inevitable trade-offs between agricultural production (e.g. carbon exports in the form of crops, fibre and 

livestock) and carbon sequestration (capture and storage) in soils. 

SOE (2011) provides a district-by-district assessment of trends in soil carbon for Australia. The assessment 

concluded the following. 

• The time since clearing is a key factor determining current trends. For example, large parts of 

Queensland are still on a declining trend because widespread clearing for agriculture was still occurring 

in the 1990s. 

• Few regions have increasing soil carbon stores. 

• Regions with intensifying systems of land use (e.g. northern Tasmania) have decreasing stores. 

• Most regions with a projected drying climate have declining trends. 

• The savannah landscapes of northern Australia have significant potential for increasing soil carbon 

stores, but this requires changes in grazing pressures and fire regimes. 

• Some of the extensive cropping lands in southern Australia with weathered and naturally infertile soils 

have not experienced as large a loss of soil carbon since clearing (e.g. generally a 30–70 percent loss and 

sometimes

New Zealand 

The importance of soil erosion to the soil carbon balance of New Zealand was noted earlier. Significant 

progress has also been made in estimating the carbon balance of different land uses. Trotter et al. (2004) 

estimated that New Zealand’s major vegetation types combined to make the total land area a small net carbon 

source (Table 15.3). A similar conclusion was reported by Tate et al. (2005). 

As indicated in Table 15.3, ‘improved’ grassland is New Zealand’s most widespread land use but there are 

large differences in carbon exchange rates within this land type. For example, detailed measurements by 

Mudge et al. (2011) over two years demonstrated that a dairy farm with mineral soil was a net C sink (year 1 = 

590 ± 560 kg C ha -1 yr -1 ; year 2 = 900 ± 560 kg C ha -1 yr -1 ) while Nieeven et al. (2005) and Campbell et al. (2015), 

using similar eddy covariance methods, demonstrated that other dairy farms with drained peat soil were net 

C sources (-1061 ± 500 kg C ha -1 yr -1 and 2940 C ha -1 yr -1 respectively). 

Vegetation type 

Area 

(million 

ha) 

Net 

carbon 

exchange 

(T g y-1) 

Source or 

sink 

Native forest 5.8 -8 Source 

Planted forest 1.6 +5 Sink 

Scrubland 3.7 -2 Source 

Native (tussock) grassland 4.3 +7 Sink 

‘Improved’ grassland (ryegrass and white clover) 6.7 -6 Source 

Unimproved grassland (species other than ryegrass and white clover) 3.4 +3 Sink 

Total for New Zealand 25.5 -1 Source 

Table 15.3 Estimated annual land–atmosphere (net) carbon (C) exchange rate for New Zealand’s major vegetation types. 

Source: Trotter et al., 2004. 

Schipper et al. (2014) reported the results of repeated sampling of 148 soil profiles across New Zealand over 

a 20-40 year period under improved pasture grazed by dairy cattle and dry stock (e.g. beef cattle and sheep). 

For soils on flat land, C stock of the uppermost 0.3 m depth decreased significantly over time (by 5 ± 21 tonnes 

C ha-1, n = 125 profiles). For silandic Andosols and Gleysols, C stock of the uppermost 0.3 m depth decreased 

significantly over time (by 14 ± 20 tonnes C ha -1 for silandic Andosols (mean ± standard deviation, n = 32), and 

by 8 ± 14 tonnes C ha -1 for Gleysols (n = 25 profiles). Soils of these groups had the highest (initial) C stocks (178 ± 

31 tonnes C ha -1 for the uppermost 0.3 m depth of silandic Andosols, 101 ± 28 tonnes C ha -1 for the Gleysols and 

96 ± 28 tonnes C ha -1 for the other soils, n = 91 profiles). For soils in hill country, C stock of the uppermost 0.3 m 

depth increased significantly over time (by 14 ± 22 tonnes C ha-1, n = 23 profiles). 

There have been some studies of potential causal factors of temporal change in soil C stock in New Zealand, 

including fertility and fertiliser application, irrigation and erosion. 

• Fertility and fertiliser application: Phosphorus (P) fertiliser application and P and nitrogen fertility 

status have not been found to account for the C stock trends of lowland and hill country soils beneath 

grazed pasture (Dodd and Mackay, 2011; Schipper et al., 2011, 2013; Parfitt et al., 2014). 

• Irrigation: The carbon stocks in some irrigated soils used for grazing over many decades have decreased 

compared to those receiving only rainfall (Kelliher et al., 2012). This may have been caused by greater soil 

respiration rates in the irrigated system, although other mechanisms are possible (see Kelliher, Curtin 

and Condron, 2013, Schipper et al., 2013, Condron et al., 2014). 


489 


• Erosion: As noted earlier, sheet erosion by water and landslides are important factors affecting carbon 

stocks in New Zealand. In one catchment, these processes each accounted for a loss of 0.5 tonnes C ha -1 

yr -1 (Page et al., 2004). On lower slopes, the soil may be lost or re-distributed or both. One study showed 

the C stock in erosion scars increased from 10 to 80 tonnes C ha -1 within 70years in the uppermost 0.2 

m depth of soil (Parfitt et al., 2013). However, the carbon stocks of soils forming on erosion scars are 

unlikely to return to more than ~80 percent of the pre-landslide amount because the newly developing 

soils are relatively shallow and drought-prone in summer (Rosser and Ross, 2011). 

Studies of current C stocks and the potential saturation value for the soils of New Zealand (Beare et al., 

2014) suggest there is an opportunity to increase the C stock of pastoral soils by increasing the C input rate, 

including deeper plant roots (e.g. Carter and Gregorich, 2010). However, organic matter stored deeply in soils 

is poorly understood (Rumpel and Kögel-Knabner, 2011) and determining the contribution of roots to soil C 

may not be straightforward (e.g. Dodd and Mackay, 2011). 

Pacific 

Only a few studies of soil carbon dynamics have been undertaken in the countries of the Pacific (e.g. 

Hartemink, 1998a). General statements are nonetheless regularly made about the decline in soil carbon 

associated with soil erosion, excessive cultivation and poor soil management. For example, Leslie and 

Ratukalou (2002) conclude that the small size of farm holdings in Fiji (60 percent are less than 3 ha) forces 

farmers into intensive cultivation (often mono-cropping) for high output, short-term production without (or 

with only minimal) fallow periods. Furthermore, competition for land is forcing subsistence gardens onto 

steeper slopes because of the expansion of cash cropping and grazing on the flatter lands. 

Excessive soil erosion in many sugar cane areas along with the burning of cane trash is resulting in serious 

depletion of fertility and soil loss on poorly managed farms. A 30-year study from Fiji on a series of sugarcane 

farms showed the expected decrease (15-35 percent) in topsoil organic C at the time of land clearing. However, 

a redistribution of C occurred, with increases at 30-40 cm depth (below the plough layer) and a very gradual 

increase in C at greater depth (80 cm). As a result, the overall C content of the fields increased on well-managed 

farms (Morrison and Gawander in press, Morrison, Gawander and Ram, 2005). 

In Papua New Guinea, the expansion of agriculture and uncontrolled logging are most likely causing a 

reduction in soil carbon stocks but quantitative studies to assess the magnitude of these changes are needed. 

Even on atolls, land-use change involving replacement of native vegetation by coconuts has led to significant 

declines in soil carbon, which is critical there because of the key role of organic matter in moisture retention 

(Morrison and Seru, 1985). 


The Southern Hemisphere does not have the same history of large scale industrialization as the Northern 

Hemisphere. However, soil contamination is a significant problem mainly in relation to impurities in phosphate 

fertilizers, agricultural chemicals, mining, waste disposal, former industrial sites and nuclear testing. Reviews 

of the history of soil contamination in the region are provided by Naidu et al. (1996) and more specifically for 

Australia (Tiller, 1992; Barzi et al., 1996), New Zealand (Roberts et al., 1996), Papua New Guinea (Singh, Levett 

and Kumar, 1996) and South Pacific Islands (Morrison, Gangaiya and Koshy, 1996). Throughout the region there 

are tens of thousands of contaminated sites (SOE, 2011; Ministry for the Environment, 2010) but the scale 

of the remediation task is not clear. Australia and New Zealand have a long and effective history of working 

together to coordinate the management and remediation of soil contamination but the waste management 

problem facing small islands in the Pacific is a serious and escalating problem. 


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Fertilizers 

Impurities in fertilizers and soil amendments such as lime and gypsum can include cadmium, fluorine, lead 

and mercury. Cadmium and fluorine have been of most concern in the region and the former can move from 

soil to the edible portions of plants. In Australia and New Zealand, the historically high concentrations of 

cadmium in phosphate fertilizers resulted from the use of island sources of high cadmium phosphate rock for 

fertilizer manufacture, primarily from Nauru and Christmas Island (McLaughlin et al., 1996; Loganathan et al., 

2003). A large effort has been devoted to determining the magnitude of the problem and to implementing a 

range of control measures (Warne et al., 2007; MAF, 2011; Cavanagh, 2014). In recent years, levels of cadmium 

in fertilisers have been reduced, and farming systems have been modified to manage the problem and mitigate 

future risk. For example, in New Zealand a tiered system for fertilizer management has been established. 

However, large areas of land that received heavy applications of superphosphate over decades now have 

elevated levels of cadmium. A recent analysis by de Vries and McLaughlin (2013) concluded that the present 

cadmium inputs from fertilizer in Australia are in excess of the long-term critical loads in heavy-textured soils 

for dryland cereals and that all other systems are at low risk. In New Zealand, a recent survey has shown that 

only isolated soil samples had cadmium concentrations that exceeded the upper-tier threshold value. The 

evidence to date indicates that cadmium in New Zealand soils poses no immediate concern. 

Agricultural chemicals 

Many of the more harmful pesticides and herbicides have been banned or tightly controlled in Australia 

and New Zealand. However, some residues can persist and adversely affect the environment, notably in areas 

that were, or still are, used for growing potatoes, tomatoes, cotton, bananas and sugar cane. Copper, arsenic 

and lead are contaminants associated with orchards and market gardens. A widespread problem in both 

countries has been the thousands of former cattle and sheep-dip sites contaminated with organochlorines 

such as dichlorodiphenyltrichloroethane (DDT) and other pesticides, including arsenic-based compounds. 

Urbanisation and the construction of dwellings on or near these sites pose a serious threat to human health. 

Most of these sites have been investigated and registered. There are far fewer studies on the impact of 

agricultural chemicals in the smaller island nations of the region. 

Mining 

The history of mining in the region was outlined above. In Australia, a significant number of current or 

former mining towns are affected by soil contamination, mostly associated with tailings, mine wastes and 

pollution from ore processing (e.g. Queenstown in Tasmania, Broken Hill, Captains Flat and Wollongong 

in New South Wales, Mt Isa in Queensland). At Port Pirie, dispersed heavy metals from smelting (primarily 

lead, zinc, cadmium and copper) can be detected over thousands of square kilometres, although seriously 

contaminated areas are restricted to tens of square kilometres (Cartwright, Merry and Tiller, 1976). 

In Papua New Guinea riverine disposal of processing residues, waste rock and overburden into rivers (e.g. 

Ok Tedi and Porgera) is having a large and long-term environmental impact on soils. Since 1984, the 1 000 

km long Fly River system has received about 66 million tonnes yr -1 of mining waste from the Ok Tedi coppergold-porphyry 

mine and this has caused widespread contamination and altered hydrological regimes (Bolton, 

2009; Campbell, 2011). Elevated levels of copper, zinc, cadmium and lead occur in the sediments that have 

been deposited in the alluvial systems of the Fly River but the longer-term impacts on ecosystems and human 

health are not clear. 


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Waste disposal 

The management of waste is now tightly regulated and managed in Australia and New Zealand. However, 

waste management is a major contemporary problem for the small island nations of the Pacific. Morrison and 

Munro (1999) provide an overview of the problem and broader issues of soil management on atoll islands (see 

case study below, Section 15.6). 

Industrial sites 

Even though they are less industrialised than many countries in the Northern Hemisphere, Australia and 

New Zealand have a legacy of contaminated former industrial sites. Most are point-scale but in some cases 

more widespread contamination has occurred. 

The legacy of nuclear testing 

Morrison, Gangaiya and Koshy (1996) review the legacy of nuclear testing in the Pacific and the following 

draws heavily on their account. Nuclear testing occurred throughout the 1940s and 1950s in the Marshall 

Islands resulting in serious contamination and impacts on resident populations. Sites in the Marshall Islands 

have been monitored for over 40 years and the detailed investigations include impacts on soils. Some areas 

will be contaminated forever. In undisturbed areas, most of the radioactivity (>80 percent) is in the upper 0.15 

m of the soil profile. The terrestrial food chain is still the most significant potential exposure pathway, but in 

some locations there may also be health problems from the intake of resuspended radioactive soil particles. 

Access to several islands is still restricted because of contamination. 

Nuclear testing also occurred in the 1950s and 1960s in Kiritimati (Christmas Island). Radioactivity 

concentrations in soil are consistent with global fall-out levels for a low rainfall equatorial area, and no site on 

the island presents a risk to the health of the local population or requires any restriction on land use. 

Nuclear testing in French Polynesia occurred from the 1960s to the 1990s and it has been a major 

international environmental issue. Morrison, Gangaiya and Koshy (1996) state that there is no doubt that 

the atmospheric testing did lead to significant contamination, but the impact of the underground tests is 

more difficult to assess. A small number of official investigations has been permitted on Mururoa, but these 

have been limited in extent and some areas have been excluded from investigations. The atoll is under French 

military control with imported food and restricted use of local resources so the direct impact of contaminants 

is limited. Morrison, Gangaiya and Koshy (1996) conclude that the impact on surrounding islands which are 

occupied by Polynesians living a traditional lifestyle is expected to be small, but in the absence of detailed 

investigations, this cannot be confirmed. 

Nuclear testing in Australia occurred during the 1950s and 1960s at Maralinga and Emu Field in South 

Australia and on the Montebello Islands near the coast of Western Australia. Rehabilitation of the contaminated 

Maralinga test area has taken decades with a major effort concluding in 2003 (DEST, 2003). Return of the 

Maralinga test area to its traditional owners was completed in November 2014. 


Soil acidification is an insidious and widespread problem in many parts of the region. If not corrected, the 

slow process can continue until the soil is irreparably damaged. The severity and extent of acidification are 

increasing in many areas due to inadequate treatment, intensification of land management, or both. Soil 

acidification is of greatest concern in situations where: (i) agricultural practices increase soil acidity (e.g. use 

of high-analysis nitrogen fertilisers or large rates of product removal); (ii) the soil has a low capacity to buffer 

the decrease in pH (e.g. infertile, light-textured soils); or (iii) the soil already has a low pH. 


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The main onsite effects of acidification include: loss or changes in soil biota involved in nitrification, 

accelerated leaching of plant nutrients, and induced nutrient deficiencies or toxicities. There may also be 

breakdown and subsequent loss of clay materials from the soil, the development of subsoil acidity, and 

reduced net primary productivity and carbon sequestration. 

The potential offsite effects on waterways include mobilisation of heavy metals, acidification, increased 

siltation and eutrophication. The process of acidification considered in this assessment is distinct from 

that associated with acid sulphate soils. Such soils occur primarily in coastal settings or with mines rich in 

sulphides. In these areas, soils naturally contain metal sulphides that severely acidify when oxidised. 

Australia 

Acidification is known to affect about half of Australia’s agriculturally productive soils. In 2001, the estimated 

annual value of lost agricultural production due to soil acidity was AUD $1 585 billion, about eight times the 

estimated cost of soil salinity at that time. More recent studies have confirmed the scale of the problem (e.g. 

Lockwood et al., 2003, SOE, 2011, DAFWA, 2013). 

Soil acidification is widespread in the extensive farming lands of southern Australia and the rates of 

lime application are well short of those needed to arrest the problem (see below). Acidification is common 

in intensive systems of land use (tropical horticulture, sugar cane, dairying). In some regions, acidification 

is limiting biomass production but the degree of restriction is difficult to estimate. Trends in the tropical 

savannahs are uncertain. If acidification is occurring there, it will be a difficult problem to solve (Noble, Cannon 

and Muller, 1997; Noble et al., 2002). Carbon losses are most likely occurring across regions in poor condition, 

and unabated soil acidification will be a major constraint on storing carbon in soils in the future. 

Ultimately, soil acidification restricts options for land management because acid-sensitive crops and 

pastures cannot be grown. It is relatively straightforward to reverse short-term soil acidification through the 

application of lime. However, it is much harder to reverse the problem if the acidification has advanced deeper 

into the soil profile, because incorporating lime at depth is prohibitively expensive. Prevention rather than 

cure is essential. While rates of lime application appear to be increasing (due to active extension programs), 

they still fall far short of what is needed to arrest the problem. The case study of southwest Western Australia 

(see below, Section 15.6) provides more details. The rates of lime application are still much lower than what is 

needed to avoid irreparable damage. In South Australia the average quantity of lime sold annually over the past 

decade (113 000 tonnes) is only 53 percent of that needed to balance the estimated annual soil acidification 

rate (SOE, 2011). 

New Zealand 

The acidification of legume-based pastures as a result of nitrate leaching and nutrient transfer/removal 

is of particular concern in New Zealand (de Klein, Monaghan and Sinclair, 1997). In many productive lowland 

pastures soil acidification is overcome through regular liming. However, amelioration of soil acidity by liming 

is not considered to be feasible in most hill country areas due to the high cost of aerial application (Bolan 

and Hedley, 2003; Moir and Moot, 2010). These soils are not cultivated and this enhances the risk for subsoil 

acidification. As a result acidification is affecting a significant portion of hill-country soils in New Zealand, 

which has the potential to significantly affect the capacity of many soils to sustain plant and animal production. 


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Papua New Guinea, Fiji and other tropical countries 

Soil acidification is a problem in areas used for sugar cane in Papua New Guinea (e.g. Hartemink, 1998b) and 

in agricultural areas in Fiji (see Section 15.5.6). It may also be significant in tropical and subtropical areas that 

have been cleared but there have been few investigations. 


Dryland salinity 

Dryland salinity is widespread across many parts of southern and eastern Australia. A large proportion of 

Australia’s agriculture is undertaken in areas with a rainfall of 450–800 mm yr -1 . In their natural condition, 

these landscapes had minimal deep drainage (generally less than 20 mm yr -1 ), and natural stores of salt brought 

in by rain and dust had accumulated in the soil in many regions. The removal of native vegetation changed the 

hydrological cycle, because trees and shrubs intercept significant quantities of rain (typically 10–20 percent of 

rainfall fails to reach the soil surface). When vegetation is removed, more water either infiltrates or runs off the 

surface. If the original vegetation has been replaced by more shallow-rooted species that use less water (e.g. 

annual crops and pastures), even more water passes through the soil. This may lead to rising groundwater 

levels and, in some cases, dryland salinity. 

Dryland salinity has been one of Australia’s most costly forms of land degradation. The comprehensive 

assessment by NLWRA (2001b) concluded that, assuming no changes in water imbalance, areas with dryland 

salinity were expected to increase from 5.7 million ha to 17 million ha by 2050. In many regions, the initial and 

sometimes primary impact of the change in hydrological regime is an increase in the salinity of streams and 

rivers (SOE, 2011). However, in low relief landscapes, large areas can be salinized and this dramatically reduces 

the options for land use. 

Between 2001 and 2009, the Millennium Drought (van Dijk et al., 2013) across southern and eastern Australia 

appears to have slowed the spread of dryland salinity in these regions but the outlook is still problematic. 

Current projections of climate change are for a drying of southern Australia which should lead to a lessening of 

the problem. However, the long-term outlook for more recently cleared land in the northern Murray–Darling 

Basin and central Queensland is unclear. Large areas are yet to reach a new hydrological equilibrium after 

clearing. However, close surveillance of groundwater systems is essential, particularly in regions that returned 

to wetter conditions in the last five years. The case study below for southwest Western Australia suggests that 

the expansion of dryland salinity may be experiencing a temporary lull in many districts. Dryland salinity is not 

a significant problem in other parts of the Southwest Pacific Region. 

Salinity in irrigation areas 

Salinity has been a major problem in the irrigation districts of Australia, particularly in the south of the 

Murray-Darling Basin where most irrigation is concentrated. Salinity developed in the early 1900s soon after 

the first schemes were completed. The scale of the problem started to be fully recognized in the 1970s and 

by 2000 it was viewed alongside dryland salinity as the country’s highest priority natural resource problem. 

Major investments in infrastructure, large-scale salt interception schemes, institutional reform, substantial 

improvements in water-use efficiency and the Millennium Drought all contributed to a mitigation of the 

problem (e.g. MDBA, 2010). SOE (2011) were equivocal in their assessment of whether salinity was still a 

major problem. However, the conclusions above for dryland salinity apply equally to irrigation salinity. New 

irrigation development is occurring in Tasmania where landscapes have stores of salts in some settings. 

Irrigation developments in northern Australia and the east coast are being explored at present but salinity 

risks in these areas tend to be less. 


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Saltwater intrusion 

Saltwater intrusion is a critical issue for the management of freshwater groundwater systems on most 

of the 1 000 or more inhabited atoll islands in the region. This issue is considered in the case study below. 

Throughout the region, increasing groundwater extraction (e.g. for irrigation), periods of below-average 

rainfall, urban development and sea-level rise have increased the risk of saltwater intrusion in coastal districts. 

In Australia, Werner (2010) and Ivkovic et al. (2012) provide an initial assessment of the problem. There is a good 

awareness of salt water intrusion risks in New Zealand and only a small number of actual salt water intrusion 

problems have occurred. Most have been into shallow unconfined aquifers and the problems have been shortlived 

and adequately managed by changes to groundwater abstraction (Callander, Lough and Steffens, 2011). 

Sodicity 

Australia has the largest extent of naturally sodic soils of any continent (FAO, 1988) but they are relatively 

rare in other parts of the region. There is a large scientific literature on the management and amelioration 

of sodicity in Australia and reviews are provided by Loveday and Bridge (1983), Summer and Naidu (1998) and 

Rengasamy (2002, 2006). Sodic soils are difficult to manage because of their poor soil–water and soil–air 

relations. Swelling and dispersion of sodic aggregates reduces the porosity and permeability of soils and 

increases soil strength even at low suction (e.g. high water content). Sodic soils have a narrow non-limiting 

water range (Letey, 1985) and they are typically either too wet immediately after rain or too dry within a few 

days for optimal plant growth. The dispersive clays cause surface crust and seal formation, poor internal 

drainage, and in some settings, tunnel erosion. 

Rengasamy (2002) estimated that more than 60 percent of the 20 million ha of cropping soils in Australia 

are sodic and that the actual yield of grains on these soils is often less than half of the potential yield. However, 

sodicity is often associated with other subsoil constraints to root growth including alkalinity, boron toxicity 

and salinity so remediation is neither easy nor economic in many cases. 

There are few studies on the dynamics of sodicity. Apart from irrigation systems where water supplies 

contain appreciable sodium, there is limited evidence to indicate that sodicity is increasing or decreasing 

in either severity or extent or both. In irrigation systems with appreciable sodium, salt loads are generally 

managed effectively. However, land-based disposal of effluent can be constrained by increasing sodicity and 

salt loads (Toze, 2006; Balks, Bond and Smith, 1998; Bond, 1998). 

Irrigation with waters containing appreciable quantities of potassium is a common occurrence, particularly 

with waste water. This monovalent cation has a similar effect to sodium, causing clay dispersion and reduced 

permeability. Smith, Oster and Sposito, (2014) state that the deleterious effect of potassium is estimated to 

be about one-third of that of sodium. 


Soil biology has been an active field of research in the region for many decades (e.g. CSIRO, 1983; Pankhurst 

et al., 1994; Abbott and Murphy, 2007). Much of this has had a focus on agriculture and forestry. However, there 

has also been a concerted effort to understand the evolution, distribution and status of biodiversity. Woodman 

et al. (2008) provide some preliminary assessments for Australia and outline strategies for developing a longerterm 

system for assessing soil biodiversity. Several significant assessments are underway in the region using 

molecular biological techniques. The Biome of Australia Soil Environments (BASE) is collecting samples to 

create a large-scale genomic database of soil biomes across Australia. Until these are completed, it is difficult 

to provide a definitive assessment of the loss of biodiversity in the region. However, some general inferences 

can be drawn. 


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• Above-ground biodiversity is much better characterized than soil biodiversity and there are several 

megadiverse districts and countries in the region. 

• Areas with high above-ground biodiversity are likely to be similarly diverse below ground if their 

environments were relatively stable throughout the Pleistocene (e.g. the moist forests of southwest 

Western Australia, rainforests on the eastern escarpment of Australia, intermediate elevations in 

Papua New Guinea, southern districts in the South Island of New Zealand). Some ancient landscapes 

were submerged during this period (e.g. New Caledonia) and this most likely had an impact on soil 

biodiversity present today. 

• A range of pressures and stressors directly affect soil biota and some have already been discussed (e.g. 

loss of soil carbon, acidification, physical disruption through cultivation or other means, intensification 

of fire regimes). Large areas in the region have experienced these pressures and stressors. As a 

consequence, they are most likely experiencing a loss of soil biodiversity. 


No comprehensive survey or monitoring of waterlogging have been undertaken at the district or national 

level in the region. Waterlogging is a significant constraint on agricultural production and extensive drainage 

schemes were installed during the twentieth century, particularly in low lying alluvial areas in New Zealand 

and eastern Australia. Texture-contrast soils with impermeable B-horizons are widespread in the pasture and 

cropping lands of southern Australia. Waterlogging is a major limiting factor of crop production in southwestern 

Victoria (McDonald and Gardner, 1987). Raised beds are sometimes used to minimize the impact of 

water logging and enhance crop production. 


Nutrient imbalances are widespread throughout the more intensively managed landscapes of the region. 

Some of these have already been outlined in relation to carbon balances and acidification (see below as well). 

The focus here is on systems of land use where nutrient mining, depletion or accumulation may be occurring. 

Nutrient mining refers to situations where there is a large removal of nutrients with minimal additions. 

In Australia, this has occurred in some extensive, low-input farming systems. For example, Dalal and Mayer 

(1986) document the decline in soil fertility over 70 years in areas used for dryland cropping in Queensland. This 

extensive low-input system relied on the natural fertility of its predominantly heavy clay soils (Vertisols) but it 

now requires fertilizer inputs to offset nutrient exports in harvested products. 

Nutrient decline is occurring in other parts of the region although there are few reliable surveys and 

monitoring systems except in New Zealand. The shortening of rotations in the shifting agricultural systems 

of Melanesia caused by increased population is most likely causing nutrient decline but minimal evidence is 

available. Nutrient decline is also likely to be occurring on marginal lands that are degrading due to processes 

such as acidification and erosion. The magnitude of the nutrient loss has been documented in a few districts, 

particularly where sediments and nutrients have a major environmental impact. 

Nutrient accumulation is a more recent phenomenon in the region and the case study on New Zealand 

(see below) explores several aspects. In most other Australian farming systems, fertilizer use has increased 

and most nutrient imbalances are managed because of the economic consequences of over- or under-use. 

In Australia, the use of nitrogen fertilizers has more than doubled in the last 25 years but application rates 

are still moderate compared to more intensively managed systems in China, Europe and the United States. 

Nutrient accumulation has occurred across southern Australia and elsewhere. In a few cases, environmental 

impacts are significant (e.g. the case of the Great Barrier Reef mentioned earlier and the Peel-Harvey system 

in Western Australia (e.g. Ruprecht, Vitale and Weaver, 2013). 


496 


Phosphorus is naturally deficient across large parts of the region. Despite this, in Australia phosphate 

fertiliser use is relatively inefficient (McLaughlin, Fillery and Till, 1990; Weaver and Wong, 2011). Most 

phosphorus is applied in the higher rainfall areas of southern Australia, and around 40 percent is applied to 

pastures. Nationally, approximately 20 percent of the phosphorus applied as fertiliser is extracted in food 

and fibre products for export, and about 5 percent is consumed domestically. The remaining 75 percent 

of the phosphorus applied in Australian agriculture accumulates in the soil, and some of this is lost to the 

environment, with detrimental impacts on waterways (Simpson et al., 2011; McLaughlin et al., 2011). The 

accumulation of phosphorus is greatest in soils across southern Australia (e.g. Weaver and Summers, 2013). 

Apart from the environmental risks caused by this accumulation, the inefficiency has an economic cost that 

will increase if fertiliser prices rise. In contrast, many grazing systems in northern Australia have pastures and 

animal production systems that are limited by deficits in phosphorus availability (McIvor, Guppy and Probert, 

2011). 

Nutrient imbalance has been a persistent issue in soils used for sugar cane in Fiji. Production began in the 

1880s and fertiliser use increased after the Second World War. Up to the 1980s, the use of particular fertilizers 

resulted in excess inputs of N and less than appropriate inputs of P and K (Morrison et al., 1985, 2005). After 

significant debate and review, blended fertilisers were introduced in the early 1990s and this has led to a more 

balanced input of N, P and K. However, topsoil Ca and Mg contents have dropped dramatically and this has 

been attributed to a continuous removal of Ca and Mg in harvested cane and by erosion. In most years, less 

than replacement quantities of these elements are being added in fertilisers. Another contributing factor has 

been a dramatic decrease in liming on sugarcane farms. 


Until recently, heavy machines such as tractors, harvesters and trucks were driven over most agricultural 

areas in the region leading to widespread soil compaction (Tullberg, 2010). Damage was greatest when the 

soil was wet. Some of the compaction can be undone through cultivation, although it is common for plough 

pans to develop just below the depth of cultivation. The distribution of pressure under a heavy vehicle also 

results in a zone of compaction halfway between the wheels, usually at a depth of around 0.5 metres. This 

type of compaction is difficult to remove. Heavy animals can also compact wet soil, leading to a decline in 

pasture production. Most of the damage occurs in the upper part of the soil profile. 

In Australia, the extent of soil compaction due to wheel traffic and its agronomic consequences have not 

been investigated in detail; however, it is thought to be very widespread (Chan et al., 2006; McGarry, Sharp and 

Bray, 1999; Tullberg, Yule and McGarry, 2007). A number of studies have documented the extent and severity 

of compaction under different systems of land use including: irrigated agriculture (McGarry, 1990), cotton 

(Braunack and Johnston, 2014), dryland cropping (Bridge and Bell, 1994), sugar cane (Braunack, Arvidsson 

and Hakansson, 2006), and grazing and forestry (Rab, 2004). One study (Geeves et al., 1995) surveyed the 

physical and chemical properties of 78 soils under crops and grazed pastures judged to be representative for 

southern New South Wales and northern Victoria. The study concluded that soil-based constraints exist in 

surface and subsurface soil horizons and that these constraints are in some instances severe. The bulk density 

of B horizons ranged from 1.25 to 1.95 Mg m -3 with a mean of 1.58 Mg m -3 indicating that root growth will be 

restricted in many of these clay-rich horizons (Jones, 1983). 

Encouragingly, there has been a rapid uptake of controlled-traffic farming in Australia and New Zealand 

during the last 15 years. This confines compaction to the smallest possible area and has the potential to alleviate 

further damage. Tullberg, Yule and McGarry, (2007) provide a detailed account of these systems and their 

impact on soils. The improved productivity, practicality and economic viability of controlled-traffic systems 

have led to enthusiastic adoption by farmers in the region. In Western Australia, farm profit has been shown 

to increase by 50 percent through the adoption of controlled-traffic farming (Kingwell and Fuchsbichler, 2011). 


497 


In New Zealand, deterioration of soil physical condition is common in grazed pastoral systems due to 

animal treading. Sparling and Schipper (2004) in their survey on the condition of soils in New Zealand reported 

moderate compaction on a large proportion of pasture soils and that about half the sites under dairying were 

below the critical threshold of 10 percent air-filled porosity. Mixed cropping soils also had low macroporosity. 

Drewry (2006) indicated that the physical condition of the top 10-15 cm naturally recovers when animals are 

partially or completely excluded from pasture and proposed the integration of a recovery cycle when managing 

grazing systems. 


Most towns and cities in the region were established during the 19th century and they were nearly all built 

on, or adjacent to, land highly suited to horticulture and cropping. The encroachment of urban and peri-urban 

development has seen the capping of this land. For economic reasons, it is highly unlikely that these good 

quality soils will ever regain their biological function. The loss of these strategically important soils is occurring 

throughout the region and is particularly significant in Fiji, some of the small Pacific islands, New Zealand and 

Australia. Population projections (Table 15.2) indicate that the problem will intensify in coming decades. The 

rate of sealing and capping is not being monitored at present. 


15.6.1 | Case study one: Intensification of land use in New Zealand 

Significant changes in land use have occurred across New Zealand in recent decades. The impacts of 

these changes on soil resources are reasonably well understood because of a substantial commitment to 

fundamental and applied soil research within the country. This effort is integral to the current and future 

international success of New Zealand’s agriculture sector. 

An important development in the evolution of land use in New Zealand that contributed to intensification 

was the removal of all agricultural subsidies in the mid-1980s (MacLeod and Moller, 2006). New Zealand 

is unique because it is the only developed country to be largely exposed to international markets. As a 

consequence, current land management decisions are now driven directly by the response of land managers 

to market prices for agricultural products. 

The removal of subsidies triggered a decline in traditional sheep farming. Sheep numbers dropped from 

around 69 million in 1980 to 39 million by 2002 and 31 million in 2012. Plantation forests were established on 

former sheep pastures and forestry expanded from 0.85 million ha in 1980 to 1.8 million ha in 2000, with 95 

percent under private ownership. The strong international demand for dairy products caused an increase in 

the number of dairy farms. Dairy cattle increased from around 3.0 million in 1980 to 5.3 million by 2002 and 

6.4 million by 2012. Cattle stocking rates and productivity also increased (Ministry of Agriculture, Fisheries 

and Food, and Statistics New Zealand). Some of the pastures converted to plantation forests in the 1980s and 

1990s are now being converted to dairy pastures following the harvesting of mature trees, and even partly 

grown trees have been removed and replaced by pasture for dairy farming (Sparling et al., 2014). Horticulture 

has also expanded. The area of horticultural crops has increased by 40 percent in just over 10 years (HortNZ, 

2014). 

Most of the intensification of agriculture in New Zealand has occurred on the better class lands. There 

has also been the expansion of dairy farming onto soils that have previously been less intensively farmed, 

such as artificially drained Luvisols in the Southland region and irrigated stony Fluvisols and Leptosols of the 


498 


Canterbury region. There has been a six-fold increase in N fertilizer use over the past 20 years (compared to 

a two to three fold increase in Australia) along with substantial increases in use of most other agricultural 

inputs. 

Soil-based primary production is fundamental to the New Zealand economy and agricultural, horticultural 

and forest products together account for 60–65 percent of the country’s total export income. Understanding 

the pressures and changes occurring in the soils of New Zealand, especially under pasture systems, has been 

a major task for New Zealand soil scientists. The establishment of a coordinated soil monitoring capability 

(Sparling and Schipper, 2002) has been important. 

Taylor et al. (2010) provide an example of the insights gained from the monitoring program for the Waikato 

region which covers much of New Zealand’s central North Island. This monitoring program is linked to systems 

for setting regional targets for various aspects of the soil resource. About 58 percent of the region is in pastures, 

18 percent is in plantation forestry and

15.6.3 | Case study two: Soil management challenges in southwest Western Australia 

The southwest of Western Australia is dominated by ancient landscapes and widespread sandy soils that 

are strongly weathered. By world standards, the soils are infertile and have a range of physical and chemical 

constraints to plant growth. During the last one hundred years, large areas have been cleared for agriculture. 

The original perennial, deep-rooted vegetation has been replaced with shallower-rooted annual crops and 

pastures. Despite significant soil constraints, extensive cropping and pasture systems have developed, 

benefitting from the existence of regular winter rains. The farming systems generally operate with low inputs 

of fertilizer, farm chemicals and soil ameliorants. Despite this, the region generates a large proportion of 

Australia’s agricultural exports. 

The serious soil management challenges of the region are reasonably well understood and it is clear that 

the pressures of climate change and degradation of soil resources are combining to threaten the viability of 

many agricultural businesses. This case study focuses on just a few of the soil management challenges. The 

account is based almost exclusively on the comprehensive report card for sustainable natural resource use for 

agriculture (DAFWA, 2013). 

The climate of southwest Western Australia is changing and over recent decades mean temperatures have 

risen and annual rainfall has declined (Figures 15.3(a) and (b)). The pattern of rainfall is also changing with 

declines in autumn and winter rainfall and increases in spring and summer rainfall. Predictions indicate that 

these trends will continue and in the short term, year-to-year climate variability may be more important for 

agriculture than the longer term trends. 


500 


Figure 15.3 (a) Trends in winter rainfall in south-western Australia for the period 1900–2012. Source: Australian Bureau of 

Meteorology 1 . 

The 15-year running average is shown by the black line. (b) Annual mean temperature anomaly time series map for south-western Australia (1910–2012), 

using a baseline annual temperature (1961–1990) of 16.3 °C. The 15-year running average is shown by the black line. 

Soil acidification 

Soil acidity in Western Australia is estimated to cost broad-acre agriculture AUD $498 million per year 

(Herbert, 2009) or about 9 percent of the average annual crop. It is one of the few soil constraints (particularly 

subsurface constraints) that can be treated with appropriate management. 

Between 2005 and 2012 a total of 161 000 samples was collected from over 93 000 sites to determine 

soil pH (determined in calcium chloride solution - pHCa) status and trend. Figure 15.4 shows the proportion 

of samples below the nominated targets for Western Australia for the surface layer (0–10 cm) of pHCa 5.5 

(desired target) and pHCa 5.0 (critical threshold). Soil acidity is widespread and extreme in many areas of the 

southwest of Western Australia, particularly in sandy soils. Surface soil pH can be increased to above target 

(pHCa 5.5) over significant areas with the application of 1 to 3 tonnes ha -1 of good quality lime. This will cost 

from AUD $50 to $150 per ha. If soil surface pH can be raised and maintained above the target this will ensure 

that management of subsurface acidity will be achieved over time. 

There is general recognition that lime use needs to increase and the trend is positive (Figure 15.5). However, 

current agricultural lime sales are still only 40 percent of the estimated annual amount required to reach the 

recommended targets over the next 10years (Gazey et al., 2013). 

1 http://www.bom.gov.au/ 


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Figure 15.4 Percentage of 

sites sampled (2005–12) 

with soil pH at 0–10 

cm depth below the 

established target of pHCa 

5.5 (left) and the critical 

pHCa 5.0 (right). Grey 

indicates native vegetation 

and reserves. Source: 

Gazey, Andrew and Griffin, 

2013. 


502 


Figure 15.5 Agricultural lime sales 2005–12 in the south-west of Western Australia based on data for 85–90 percent of the market. 

Other soil management challenges 

A range of other serious threats to soil function are prevalent in southwest Western Australia. Dryland 

salinity has been accelerated and enlarged by widespread clearing and land-use change. The problem affects 

agricultural land, water resources, natural biodiversity and the built environment with an economic cost of 

hundreds of millions of dollars annually. Currently, more than 1 million ha are severely salt-affected (McFarlane, 

George and Caccetta, 2004; Furby, Caccetta and Wallace, 2010). Dryland salinity has expanded in most of this 

region since 1998, especially following episodic rainfall events (George et al., 2008a; Robertson et al., 2010). In 

areas cleared and developed for agriculture after 1960, most water tables continue to rise, despite a decline 

in annual rainfall. Dryland salinity is a potential threat to 2.8–4.5 million ha of productive agricultural land 

(George et al., 2005). The long-term extent of salinity may take decades to centuries to develop (George et al., 

2008b). The areas with the highest dryland salinity risk occur mostly in the highly productive, medium to high 

rainfall dryland agricultural areas. Management to contain or adapt to salinity is technically feasible using 

plant-based and engineering options, though recovery is economically viable in only a few areas (Simons, 

George and Raper, 2013). 

Wind erosion is a critical issue for the livestock and grain industries. Maintaining sufficient, stable soil 

cover throughout the year is the challenge and this is difficult in a variable and generally drying climate. Water 

erosion hazard during the growing season has diminished due to declining winter rains and more sustainable 


503 


management practices. Water erosion events are mainly caused by intense, localised summer storms, and 

these are likely to increase with changing climate. Maintaining sufficient soil cover to prevent water erosion is 

not always possible in grazing systems. 

Water repellence is widespread on the sandy soils of the region. The annual average opportunity cost of 

lost agricultural production is estimated to be AUD $251 million. Water repellence also increases the risk of 

wind and water erosion, off-site nutrient transport and possibly soil acidification through increased nitrate 

leaching. The extent and severity of water repellence appears to be increasing as cropping increases together 

with early sowing, minimum tillage and reduced break of season rainfall. 

A warming and generally drying climate associated with lower organic matter inputs could limit future 

sequestration of organic carbon in soils used for livestock and grains production. Farming systems that increase 

biomass production and minimise processes such as wind and water erosion are required to maintain and 

improve soil carbon levels. Soil compaction is suspected to be a major constraint for all agricultural industries 

in the region but investigations are needed to quantify its extent and severity. Finally, the nutrient status of 

soils has increased in agricultural systems. On average, pasture soils and arable soils contained 1.3 times and 

1.6 times respectively as much phosphorus (P) as is required for optimum production. 

15.6.2 | Case study three: Atoll Islands in the Pacific 

The fifth assessment report of the IPCC (AR 5) concluded that current and future climate-related drivers 

of risk for small islands during the 21 st century include sea level rise, cyclones, increasing air and sea surface 

temperatures, and changing rainfall patterns. This confirmed previous IPCC assessments and highlighted the 

vulnerability of small islands to multiple stressors, both climate and non-climate (Nurse et al., 2014). Limited 

water and soil resources are additional stressors that contribute to the vulnerability of small islands, and 

especially atoll islands, in the Southwest Pacific region. 

The following is drawn from Morrison’s (1990, 2011) reviews of soils on atoll islands. The atoll islands are in 

essence reefs of variable thickness that have built up by corals resting on a volcanic base. There is a variety of 

atoll forms but most commonly they occur as a slightly emerged calcareous reef surrounding a lagoon. Low 

atolls are usually less than 5 m above sea level. Raised atolls (raised coral platforms) can be higher (e.g. Nauru 

and Niue). Many atolls have deposits of ash or pumice from nearby volcanic activity 

The sandy textured soils on most atoll islands are highly calcareous and very permeable. They usually show 

minimal profile development. The soils are heavily dependent on organic matter to ensure the retention 

and availability of water and nutrients. Organic carbon can range from 1 to 20 percent in the surface layers 

depending on the vegetation and position in the landscape. There is minimal organic carbon deeper in 

the profile. The calcareous mineralogy and associated alkalinity cause a range of plant nutrition problems 

(e.g. deficiencies of K and micronutrients). The soils are prone to periods of water stress unless irrigated. In 

summary, the soils are infertile and poorly suited to intensive agriculture. 

Population growth (Table 15.2) and increasing urbanization combined with the decline of traditional 

agriculture has left many of the small island nations largely dependent on imported processed food. This 

dependency on processed food is a significant factor contributing to poor health associated with the rise of 

chronic diseases like hypertension and diabetes. Local production of fresh food is therefore a necessity on atoll 

islands and Deenik and Yost (2006) provide a case study for the Marshall Islands. 

The fertility of soils on atoll islands can be improved through the incorporation of composts and organic 

mulches. This was a central feature of traditional systems. Mechanised mulching and composting systems 

are being developed on several islands along with various irrigation systems that aim to improve water use 

efficiency. However, various nutritional and soil management issues have to be resolved, particularly in 

relation to the suitability of different feed stocks and the unusual soil chemistry of some islands. 


504 


The management of water is also fundamental because freshwater reserves on most atoll islands are 

under significant stress due to excessive groundwater extraction, contamination (e.g. from waste dumps) 

and saltwater intrusion. Finally, the farming systems on low atoll islands are vulnerable to salinization due to 

either poor quality groundwater associated with saltwater intrusion or infiltration of seawater after storm 

surges and flooding. 

At present, there is very limited soil science expertise available to help develop more productive and 

sustainable agricultural systems on atoll islands. White, Falkland and Fatai (2009), in reviewing the 

management of freshwater lenses on small Pacific islands, conclude that they are some of the most vulnerable 

aquifer systems in the world. This conclusion applies equally to the soil and food production systems on atoll 

islands. 

15.6.4 | Case study four: DustWatch – an integrated response to wind erosion in Australia 

Wind erosion has environmental impacts at the source where soils are eroded (onsite wind erosion), and 

much greater economic and human health impacts downwind from the source where air quality is reduced 

(offsite wind erosion). Wind erosion has been a major problem in Australia and it continues to be significant 

although incidence and severity have declined since the 1940s. The good technical understanding of wind 

erosion in Australia has relied heavily on the development of DustWatch – a collaborative effort to monitor 

wind erosion across the country. This case study summarizes the status of wind erosion in Australia and 

provides a brief introduction to DustWatch. 

Wind erosion trends for Australia 

Climate is by far the strongest determinant of wind erosion. Land management can either moderate or 

accelerate wind erosion rates. Unravelling these two influences has been difficult, but between 2001 and 2009 

the Millennium Drought – the worst on record (Van Dijk et al., 2013) – provided an opportunity to gauge the 

effectiveness of improvements in land management that had occurred in recent decades across the country. 

Historical accounts indicate that wind erosion was very active during the drought periods of the late 19th 

and early 20th centuries (e.g. Ratclife, 1938). While these anecdotal reports present dramatic images of huge 

dust storms engulfing rural towns, and sand drifts burying fence lines and blocking rural roads, until recently 

it had not been unequivocally established whether the ‘dust bowl years’ of the 1940s were due to extreme 

drought, poor land management or both. 

McTainsh et al. (2011) have analysed wind erosion activity during the 1940s and 2000s. They used archived 

meteorological data to calculate the dust storm index (DSI) for both periods (McTainsh, 1998, McTainsh et 

al., 2007). The DSI, which provides a measure of the frequency and intensity of wind erosion activity, is the 

accepted measure of wind erosion activity in Australia (Bastin, 2008, SOE, 2011). Overall, mean onsite wind 

erosion in the 1940s was almost six times higher (mean DSI = 11.4) than in the 2000s (mean DSI = 2.0), and the 

mean maximum DSI for the 1940s was four times that of the 2000s (SOE, 2011). 

There were also significant differences between districts. The decrease in wind erosion in the 2000s was 

much more pronounced in the east and centre of the continent. The 1940s erosion rate in central Australia 

was large, but the decrease in the 2000s was less than elsewhere. There has also been a large decrease in the 

number of dust storms reaching the coastal cities. 

O’Loingsigh et al. (2015) confirmed these findings by using Dust Event Days (DED) and a modified version of 

a published Dust Storm Index (DSI) to show that wind erosion during the World War II Drought (1937-1945) was 

up to 4.6 times greater than during the Millennium Drought. 


505 


Despite these trends, the Millennium Drought resulted in large dust storms and other wind erosion activity. 

Two extreme dust storms hit eastern Australian cities on 23 rd October 2002 and 23 rd September 2009. The ‘Red 

Dawn’ dust storm on the 22nd and 23 rd September 2009 was the largest to pass over the coastal city of Sydney 

since reliable records began in 1940 (Figure 15.6). Visibility was reduced to 400 m and the maximum hourly 

PM 10 

concentration was 15 366 μg m -3 – the highest ever recorded for Sydney and possibly for any Australian 

capital city. The Australian air quality standard of 50 μg m -3 per 24 h was massively exceeded in suburban 

Sydney (1 734 μg m-3) and 150 km to the north in the city of Newcastle (2 426 μg m-3). 

Red Dawn was caused by drought and extreme winds. The source of the red dust was primarily the sand 

plains of western New South Wales, the sand plains, riverine channels and lakes of the lower Lake Eyre Basin 

and the Channel Country of Queensland. The estimate of total suspended particulate sediment lost off the 

Australian coast for the 3 000 km long Red Dawn dust storm was 2.54 million tonnes with a plume height of 

2 500 m. This is the largest off-continent loss of soil ever reported using measured, as opposed to modelled, 

dust concentrations for Australia (Leys et al., 2011). This single event resulted in economic costs of between 

AUD $293–A$313 million with most being associated with household cleaning and associated activities. The 

dust storm also affected the state of Queensland, but costs there were not included in the study by Tozer and 

Leys (2013). 

Figure 15.6 MODIS image for 0000 23 September 2009 showing Red Dawn extending from south of Sydney to the Queensland/ 

NSW border and the PM 10 concentrations measured using Tapered Element Oscillating Microbalances (TEOM) at the same time at 

ground stations. 


506 


The DustWatch network 

Australia has a long tradition of participatory and collaborative approaches to natural resource 

management. DustWatch (Leys et al., 2008, DustWatch, 2014) fits within this broader tradition. It was 

established to fill some of the gaps in the measurement networks of wind erosion and to supplement land 

management information provided by community participants. 

Measuring and monitoring wind erosion is problematic because of the irregular nature of wind erosion 

in space and time. DustWatch aimed to build on the long record of observations collected by the Australian 

Bureau of Meteorology (BoM) by strategically adding a community network of observers and instruments 

to fill identifiable gaps in the BoM network. DustWatchers use the same basic measurement protocols as 

the BoM. In addition, satellite imagery is used wherever possible to help define source areas and dust 

paths. Remote sensing alone has not been successful due to the presence of cloud and the low resolution 

of geostationary satellites and the low temporal resolution of satellites with better spatial resolution. 

DustWatch has successfully mixed the formal reporting of wind erosion activity from the BoM with informal 

community networks, instrumented sites and ad hoc satellite data to provide an understanding of wind 

erosion in Australia. The major outcomes from DustWatch have been: 

• better reporting of the extent and severity of wind erosion across Australia at national to regional scales 

• greater dialogue between the scientists and the community 

• increased awareness of the impact of wind erosion on the environment 

• documentation of local community knowledge on why and where wind erosion is occurring 

DustWatch is important because it has provided new data sources when other formal data sources are 

inadequate. This includes valuable land management information from local landholders. The spatial and 

temporal data are also provided at the correct scales for testing physical models and remote sensing products. 

This ultimately has provided the capability for understanding how wind erosion processes are influenced by 

climate and land management. 

15.7 | Conclusions 

Based on the above finding, a provisional assessment is made of the status and trend of the 10 soil threats in 

order of importance for the region. At the same time an indication is given of the reliability of these estimates 

(Table 15.4). 

The status of soil resources in the Southwest Pacific region is mixed. The region is a globally significant 

exporter of agricultural products and nearly all of the 24 countries rely heavily on soils for wealth generation. 

The threats to soil function in some countries are serious and require immediate action to avoid large scale 

economic costs and environmental losses. These threats to soil function combined with other pressures 

caused by increasing population and climate change are especially challenging in southwest Western Australia 

and on the atoll islands of the Pacific. It is difficult to assess some threats because of the lack of surveys and 

monitoring networks. The example of soil monitoring from New Zealand demonstrates the practical value of 

having the capability to track and respond to soil change. The intensification of land use in New Zealand and 

to a lesser extent Australia provides an indication of the soil management challenges that will dominate in 

comingyears as countries attempt to substantially increase food production within a resource constrained 

world. Poor land management practices, and especially uncontrolled logging in the low-income countries in 

the Southwest Pacific are a significant challenge to national prosperity. 


507 



function 


acidification 


Organic 

carbon change 

Nutrient 

imbalance 

Compaction 

Summary 

A widespread and 

serious problem that has 

the potential to cause 

irreversible damage 

to soils particularly in 

southern Australia, tropical 

landscapes and areas 

where product removal and 

leaching are contributing 

factors. 

Improved land 

management practices in 

Australia and New Zealand 

have reduced erosion rates 

but the problem is still 

serious in some districts. 

Unsustainable rates of soil 

loss are associated with 

logging and clearing in 

several Pacific nations 

The conversion of land 

to agricultural uses has 

generally caused large 

losses of organic carbon 

in soils. Improved land 

management practices 

have stabilized the 

situation but there is 

limited evidence for 

increasing soil carbon 

even under these more 

conservative management 

systems. 

Rapid intensification 

of agriculture in New 

Zealand and more recently 

Australia is causing 

significant environmental 

impact, particularly due 

to the large increase in 

fertilizer use and ruminant 

animals. In other districts, 

nutrient mining and 

decline is occurring due to 

insufficient replacement of 

nutrients removed through 

harvest or other losspathways. 

Limited evidence suggests 

the problem is constraining 

plant growth across large 

areas, particularly in the 

cropping and pasture lands 

of Australia and smaller 

areas in New Zealand. 

Controlled traffic and other 

improved management 

practices may have halted 

this decline in soil physical 

fertility. 



Confidence 

In 

In trend 

condition 


508 



and capping 

Contamination 

Salinization 


Loss of soil 

biodiversity 

Waterlogging 

Loss of good quality 

agricultural land due 

to urban and industrial 

expansion is an emerging 

and potentially major 

problem for all countries in 

the region. 

Most sources of soil 

contamination are now 

regulated and controlled 

although the legacy of past 

practices is significant (e.g. 

Cd in fertilizers). 

Contamination caused by 

mining and waste disposal 

is a significant issue for 

several Pacific nations. 

Salinization is a widespread 

and expensive problem in 

Australia and some atoll 

islands. After a temporary 

respite due to dry years, 

the problem may continue 

to expand and the time to 

equilibration is likely to be 

in the order of decades. 

Rates of loss were most 

likely highest during the 

expansion of agriculture, 

particularly over the last 

100 years, and this may 

have slowed. 

However, information on 

baselines and trends is 

lacking in nearly all districts 

and countries. 

Waterlogging is a 

constraint to agricultural 

production in some wet 

years but evidence on 

the extent and severity is 

lacking. Large areas were 

drained to address the 

problem, particularly in 

New Zealand and parts of 

coastal Australia. 

Table 15.4 Summary of soil threats status, trends and uncertainties in the Southwest Pacific 


509 


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of Soil Change in Antarctica 

Contributing author: Megan Balks 


520 

in Antarctica

16.1 | Antarctic soils and environment 

Antarctica has a total area of 13.9 × 106 km 2 , of which 44 890 km 2 (0.32 percent) is ice-free (Fox and Cooper, 

1994; British Antarctic Survey, 2005) with potential for soil development. Ice free areas are mainly confined 

to the Antarctic Peninsula, a few places around the perimeter of the continent and along the Transantarctic 

Mountains. The largest ice-free area (approximately 5 000 km 2 ) is the McMurdo Dry Valleys in the Ross Sea 

Region. 

Mean annual temperatures vary from near 0°C in the moister, marine influenced Antarctic Peninsula 

to about -20°C in dry, higher altitude inland areas (Campbell and Claridge, 1987). Many soils are formed on 

mixed tills with some directly formed on bedrock. Organic matter is minimal (< 0.1 percent) in drier, colder 

inland areas. However, in areas where free moisture occurs, mosses and, on the Antarctic Peninsula, higher 

plants, may grow and accumulate to form peat soil materials. Ornithogenic materials dominate soils at many 

coastal locations (e.g. Hofstee et al., 2006). Surface ages in coastal and Antarctic Peninsula regions tend 

to be predominantly Holocene, exposed by retreat of the Last Glacial Maximum ice. At higher elevations in 

the McMurdo Dry Valleys, surfaces as old as Mid-Miocene (14 Ma) have been reported (Sugden, Bentley and 

Cofaigh, 2006) indicating low erosion rates under a stable polar desert climate. Soil microclimates, driven by 

strong topographic variability, also influence soil properties (Balks et al., 2013). 

A wide variety of soils occur in the ice-free areas (Campbell and Claridge, 1987). Gelisols (Soil Survey Staff, 

2014) or Cryosols (IUSS Working Group WRB, 2014) are the predominant soils in Antarctica. Cryosols contain 

permafrost at depth and are overlain by an active layer that thaws during the summer and is frozen in winter. 

In moister coastal areas the permafrost is ice-cemented and thus ‘frozen solid’. However in some inland 

areas of the Trans-Antarctic Mountains, including the McMurdo Dry Valleys, there is not enough moisture to 

form ice-cement so soils with temperatures well below 0°C are loose and easily excavated (Bockheim, 1978; 

Campbell and Claridge, 1987). The soils range from Gelisols (Cryosols) in the Ross Sea Region, through Gelisols 

and Entisols in coastal East Antarctica, to a mixture of Gelisols, Entisols, Spodosols and Inceptisols in the 

warmer northern Antarctic Peninsula Region where permafrost is not ubiquitous (Balks et al., 2013). Due to 

limited weathering in the cold climate, many Antarctic soils are dominantly gravelly sands. Where vegetation 

is absent, a protective desert pavement usually forms at the soil surface. The focus for studies on Antarctic 

soils is not on their potential for food production, but rather on their genesis, diversity, and vulnerability to 

impacts of human activity. 

16.2 | Pressures/threats for the Antarctic soil environment 

Most of the human activities in Antarctica, including historic huts, modern research stations, and tourist 

visits, are concentrated in the relatively accessible, small, ice-free areas, on the coast, particularly in the Ross 

Sea region and Antarctic Peninsula (O’Neill et al,. in press). 

Antarctica was first recorded by three whaling ships in 1820 leading to regular whaling visits and the 

Ross expedition of 1839-1940. The ‘heroic era’ of exploration (1895–1917) included expeditions such as those 

of Borchgrevink, Scott, Shackleton, Mawson and Amundsen. Since the International Geophysical Year (1957- 

1958), greatly increased human activity has occurred with over 70 scientific research bases established, mainly 

around the Antarctic coast. Ship-based Antarctic tourism has become popular with 46 000 tourists reported 

in the 2007/08 summer and 27 700 in the 2013-2014 season (IAATO, 2014). 


521 

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Legacies of human occupation are scattered at isolated sites across Antarctica, particularly in areas close to 

the major research stations and semi-permanent field camps (Campbell, Balks and Claridge, 1993; Kennicutt 

et al., 2010; Tin et al., 2009). Impacts have included physical disturbance as a result of construction activities, 

geotechnical studies, and roading (Campbell, Balks and Claridge, 1993; Campbell, Claridge and Balks, 1994; 

Harris, 1998; Kennicutt et al., 2010; Kiernan and McConnell, 2001); local pollution from hydrocarbon spills 

(Aislabie et al., 2004; Kim, Kennicutt II and Qian, 2006; Klein et al., 2012) and from waste disposal (Claridge 

et al., 1995, Snape, Morris and Cole, 2001; Santos et al., 2005; Sheppard, Claridge and Campbell, 2000); 

introduction of alien species (Frenot et al., 2005; Chown et al., 2012; Cowan et al., 2011); and disturbance to soil 

biological communities (de Villiers, 2008; Harris, 1998; Naveen, 1996; Tin et al. 2009 and references therein). 

The amount of contaminated soil and waste has been estimated at 1–10 million m 3 (Snape, Morris and Cole, 

2001). The presence of persistent organochlorine pollutants in Antarctica has been attributed to long-range 

atmospheric transport from lower latitudes (Bargagli, 2008). 

Antarctic soils are easily disturbed and natural recovery rates are slow due to low temperatures and often 

a lack of liquid moisture (Campbell, Balks and Claridge, 1993, 1998a; Campbell et al., 1998b; Kiernan and 

McConnell, 2001; Waterhouse, 2001). Where physical disturbance removes the protective ‘active layer’ the 

underlying permafrost will melt with resulting land surface subsidence and, in drier regions, accumulation of 

salt at the soil surface (Campbell, Claridge and Balks, 1994; Waterhouse, 2001). Campbell and Claridge (1975, 

1987) recognized that older, more weathered desert pavements and associated soils were the most vulnerable 

to physical human disturbance. However disturbances on active surfaces, such as gravel beach deposits, 

aeolian sand dunes and areas where melt-water flows, have the capacity to recover (visually) relatively quickly 

(McLeod, 2012; O’Neill, Balks and López-Martínez, 2012b, 2013: O’Neill et al., 2012a). 

Fuel spills are the most common source of soil contamination and have the potential to cause the greatest 

environmental harm in and around the continent (Aislabie et al., 2004). Hydrocarbon fuel spills have been 

shown to persist in the environment for decades, with fuel perching on top of ice-cemented permafrost 

(Balks et al., 2002). When spilled on Antarctic soils, possible fates of the hydrocarbons include dispersion, 

evaporation, and biodegradation. Hydrocarbon degrading microbes are present in the Antarctic environment 

but within the Ross Sea region their effectiveness is limited by moisture and nutrient (N and P) availability 

(Aislabie et al., 2004, 2012). Hydrocarbon spills on Antarctic soils can enrich hydrocarbon-degrading bacteria 

within the indigenous microbial community (Aislabie et al., 2004, 2012; Delille et al. 2000). 

Elevated levels of metal concentrations have been reported at base sites especially in areas used for waste 

disposal or affected by emissions from incinerators or fuel spills (Claridge et al. 1995; Sheppard, Claridge and 

Campbell, 2000; Webster et al., 2003; Santos et al., 2005; Stark et al. 2008; Guerra et al., 2011). Particularly 

high metal levels have been reported at Hope Bay on the Antarctic Peninsula (Guerra et al., 2011) and at the 

Thala Valley landfill at Casey Station, East Antarctica (Stark et al., 2008). Elevated levels of methyl lead have 

been detected in soil from a former fuel storage site at Scott Base (Aislabie et al., 2004). 

Surface trampling has been shown to impact on soil nematode abundances in the McMurdo Dry Valleys 

(Ayres et al., 2008) and on arthropod abundance on the Antarctic Peninsula (Tejedo et al. 2005, 2009). 

Potential for introduction of invasive plant, insect, and microbial biota is gaining attention (Cowan et al., 2011; 

Chown et al., 2012; Greenslade and Convey, 2012). 


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in Antarctica

16.3 | Response 

All activities in Antarctica are regulated through the national administrative and legal structures of the 

countries active in the region, underpinned by the international legal obligations resulting from the Antarctic 

Treaty System. The Protocol on Environmental Protection to the Antarctic Treaty (the Madrid Protocol) was 

signed in 1991 and designates Antarctica as ‘a natural reserve devoted to peace and science’. The Madrid 

Protocol mandates the protection of Antarctic wilderness and aesthetic values and requires that before any 

activity is undertaken the possible environmental impacts are assessed. Since the ratification of the Madrid 

Protocol in 1991 environmental awareness has increased and the standard of prevention of human impacts 

undertaken by many of the Antarctic programmes, such as those operating in the McMurdo Dry Valleys, is 

now more stringent than environmental management standards in most, if not all, other regions of the planet 

(O’Neill et al., in press). 

The ice-free areas visited by humans are small, relative to the Antarctic continent as a whole, and impacts 

occur as isolated pockets amongst largely pristine Antarctic wilderness (O’Neill et al., in press). The most 

intense and long-lasting visible impacts occur around the current and former research bases, and are often 

remnants of activities in the 1950s-1970s prior to the Madrid Protocol (Campbell and Claridge, 1987; Webster et 

al., 2003; Bargagli, 2008; Kennicutt et al., 2010; O’Neill, 2013). Since the 1980s environmental accountability, 

management and awareness have increased, and the environmental footprints of stations such as Scott Base 

and McMurdo Station on Ross Island have remained static or decreased (Kennicutt et al., 2010). For example, 

there are mechanisms in place to prevent spills, remove wastes, phase out incineration, limit soil disturbance, 

and protect sites of particular cultural or environmental significance. These mechanisms are proving effective 

at preventing further damage to Antarctic soils. 

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Annex | Soil groups, 

characteristics, 

distribution and 

ecosystem services 

Coordinating Lead Authors: Maria Gerasimova and Thomas Reinsch 

Peer Reviewer: Neil McKenzie 

Contributing authors: L. Anjos, O. Batkhishig, J. Bockheim, R. Brinkman, G. Broll, P. Charzyński , M.R. 

Coulho, F.O. Nachtergaele, M. Nanzyo, S. Mantel, S.M. Pazos (†), M.H. Stolt, C. Tarnocai , T. Tóth, L.P. Wilding 

and G. Zhang. 

Status of the World’s Soil Resources | Main Report Annex | Soil groups, characteristics, 

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distribution and ecosystem services

1 | Soils with organic layers 

HISTOSOLS 1 

In most Histosols organic materials are deposited in a wetland environment to form peat. The waterlogged 

conditions, oxygen deficiency and, very often in the north, low temperatures, and acidic conditions, inhibit 

decomposition and lead to accumulation of organic matter (Kolka et al., 2012). In some Histosols the organic 

material is derived from upland forest vegetation under cool, wet high rainfall conditions. 

The soil profile features of Histosols reflect the origin of the organic material and the degree of decomposition, 

while the occurrence of permafrost is a common feature in these soils in arctic landscapes (FAO, 2014; Figure 

A 1). The soil materials in Histosols are generally dark brown to almost black reflecting the high organic matter 

content. These soils support forest, sedge and shrubby-moss types of vegetation and occupy a poorly-drained, 

level topography. However, some Histosols in the coastal areas are found on slopes or form a continuous cover 

on the terrain, such as blanket bogs. Most of these soils developed during the Holocene Epoch. The age of the 

basal peat (the peat layer just above the mineral contact) is usually five to nine thousand years old. 

Histosols are common soils in the Boreal and Arctic landscapes of the Northern hemisphere, although they 

may also occur in temperate and some tropical regions. Globally, peat lands (organic soils developed on peat) 

cover approximately 4 million km 2 (World Energy Council, 2013). However, most of the Histosols (3.5 million 

km 2 ) are found in the Northern Circumpolar Permafrost Zone where 76 percent of these soils are perennially 

frozen (Tarnocai et al., 2009). Nearly 80 percent of all Histosols occur in Russia, Canada and the United States. 

The global significance of Histosols is that they store huge amounts of organic carbon. It has been estimated 

that they represent a carbon pool of 500 billion tonnes of organic carbon (Strack, 2008). In addition, the 

present rate of annual carbon sequestration is approximately 100 million tonnes (Strack, 2008), which 

exceeds the present carbon loss from these soils due to agriculture, peat extraction and other human-made 

disturbances. Due to climate change, however, the water-saturated Histosols could be a source of greenhouse 

gases, mainly in the forms of methane (Couwenberg et al., 2010), but they also could be the source of carbon 

dioxide if these soils dry out and are affected by wildfires. 

1 The Reference Soil Group names of the World Reference Base developed by the IUSS Working Group RB (FAO, 2014) are used. Where the approximate equivalent 

name in the USDA Soil Taxonomy (Soil Survey Staff, 2014) is different, the USDA name is cited in brackets. 


528 


a 

Photo by C. Tarnocai 

b 

Photo by S. Khokhlov 

Figure A 1 (a) A Histosol profile and (b) a peatbog in East-European tundra. 


529 


2 | Soils showing a strong human influence 

ANTHROSOLS (Plagganthrepts, Haplanthrepts, some Orthents) 

Anthrosols (Figure A 2) are soils formed, altered or influenced by intense human agricultural activities. They 

are associated with long-term agricultural management in many parts of the world, especially where ancient 

civilizations were present. Technosols (see below) are also human-influenced but are connected with more 

recent human activities in industrial and urban environments resulting in the presence of artificial and manmade 

objects in the soil. 

The formation of Anthrosols is termed anthropedogenesis. This formation includes various processes induced 

by human activities in ancient agricultural systems, such as periodic irrigation and drainage, continuous buildup 

by applying transported manure or other soil materials, and long-term fertilization. These soils are often 

enriched with phosphorus and carbon, and are characterized by movement and accumulation of clay and 

clay-organic complexes, reduction and oxidation of iron-manganese oxides and even physical compaction, 

processes that occur at an accelerated rate compared with that of natural soil changes. This results in a special 

soil morphology and soil horizon development, such as the formation of a surface horizon with a high organic 

matter content, the development of compacted plough-pans and the formation of redoximorphic features, all 

of which represent the outcome of anthropedogenesis. 

Anthrosols occur widely across the globe. They appear, for example, in ancient agricultural regions under 

paddy cultivation, or in semi-arid and arid regions where irrigation and sedimentation have occurred. 

Anthrosols may also show a long-term build-up of elements from long-term manure application and 

phosphorus enrichment. Globally, the total extent of Anthrosols is estimated at more than 200 million ha, of 

which 80 percent are cultivated paddy fields. 

Anthrosols make up the most fertile agricultural land in the world and provide food as an essential 

ecosystem provisioning service. They often are an inherent part of unique agricultural systems and as such 

have a cultural function as well. 


530 


Figure A 2 (a) An Anthrosol (Plaggen) 

profile and (b) associated landscape in 

the Netherlands. 

a 

Photo by ISRIC World Soil Information 

b 

Photo by S. Mantel 


531 


TECHNOSOLS (Non soils) 

Technosols (Figure A 3) are common soils on all continents. They are dominant in urban areas, where there 

are only remnants of natural soils and where soils radically transformed by different human activities dominate 

together with ‘new soils’. The development of particular horizons and layers in such soils is not reflected in 

natural conditions of the system (Charzyński et al., 2013). Technogenic activities lead to the construction of 

artificial soil, soil sealing or extraction due to mining of materials not affected by surface processes in natural 

landscapes. The largest areas of Technosols in comparison to country total area can be found in countries with 

an extremely high percentage of urbanization such as Belgium and the United Kingdom. The largest areas 

dominated by Technosols are located within the largest mega-cities, for example the Yangtze River Delta 

Megalopolis in China (population of about 90 million); the Taiheiyō Belt in Japan, also known as Tokaido 

corridor (population of nearly 80 million); and the Great Lakes region in the United States (60 million). 

Technosols are soils of urban, industrial, traffic, mining and military areas. 

There are four main varieties of Technosols: 

• soils sealed by technic hard material (hard material created by humans in industrial processes) e.g. 

asphalt or concrete. 

• soils containing a large amount (more than 20 percent in the upper 1 m of soil) of artefacts. Artefacts 

are objects in the soil formed or strongly transformed by human activity or excavated from beneath the 

earth. Examples of artefacts are mine spoils, dredgings, rubbles, organic garbage, cinders, industrial 

dust, synthetic solids and liquids (e.g. petrol, kerosene) 

• soils with geomembranes or synthetic membranes made, for example, of polyvinyl chloride (PVC) laid 

on the surface or into the soil. 

• constructed or naturally developed shallow soils on buildings, without any contact to other soil 

material. 

The soil profile features of Technosols are usually weak. Beneath technogenic deposits occasionally a 

natural soil profile can be observed. Also original profile development may still be present in contaminated 

natural soils. 

In the urban ecosystem, soils play an essential role with their functions and ecosystem services. However, 

the ability of Technosols to provide ecosystem services differs from the services secured by natural soils 

and is often impaired (Morel et al., 2015; Stroganova et al., 1998). Technosols are more likely to contain toxic 

substances than other types of soils and should be treated with care (FAO, 2014). The benefits of Technosols 

and other urban soils are nonetheless numerous. They provide groundwater recharge for water supply, 

plant products for food supply, a medium for alternative storm-water management, sites for recreational 

activities, and buffering of temperature and humidity. They serve as a medium of retention, decomposition 

and immobilization of contaminants, and for dust entrapment to reduce dust content in the air. Technosols 

can be also considered as historical archives (Lehmann and Stahr, 2007). 


532 


a 

Photo by P. Charzyński 

Figure A 3 (a) 

A Technosol 

profile and (b) 

artefacts found in 

Technosol. 

b 

Photo by P. Charzyński 


533 


3 | Soils with limitations to root growth 

CRYOSOLS (GELISOLS) 

Due to the presence of perennially frozen conditions, Cryosols have unique processes and properties 

different from other soil groups. Cryogenic processes, which dominate the development of these soils, are 

driven by the mobility of unfrozen soil water as it migrates along the thermal gradient from warm to cold. This 

unfrozen water then moves into the frozen soil system, and feeds the ice bodies. The increase of ice volume 

and the volume increase from water to ice lead to differential frost heave. This then results in cryoturbation 

and the formation of cryogenic macro and micro soil structures (Figure A 4). 

Cryosols are the major soils in the permafrost areas of the Northern Circumpolar Arctic and Subarctic as 

well as a large part of the Boreal Region. They also occur in the ice-free areas of Antarctica and in the subalpine 

and alpine areas of the mountainous regions. They cover approximately 10.2 million km 2 in the Northern 

Circumpolar Region (Tarnocai et al., 2009) and approximately 46 thousand km 2 in the Antarctic Region 

(Tarnocai and Campbell, 2002). Globally, most Cryosols occur in Russia and Canada. Cryosols support forest 

vegetation in the Boreal and Subarctic regions and tundra vegetation in the Arctic and Alpine Regions. Most 

Cryosols in Antarctica are unvegetated. 

Cryosols, especially those affected by cryoturbation, contain large amounts of organic carbon. Cryoturbation 

moves organic materials from the surface into the subsoil where it is preserved for thousands of years because 

of the cold soil temperatures. Cryosols in the Northern Circumpolar Permafrost Zone contain approximately 

351.5 Gt of carbon in the 0-100 cm depth and 818 Gt in the 0-300 cm depth (Tarnocai et al., 2009). Due to 

climate change and the resulting thawing of these high carbon content Cryosols, they could be the source of 

greenhouse gases (carbon dioxide and methane), which would then further increase climate warming. 


534 


a 

Photo by A. Filaretova 

b 

Photo by A. Sadov 

Figure A 4 (a) A Cryosol profile and (b) associated landscape in West Siberia, Yamal Peninsula. 


535 


LEPTOSOLS (lithic sub-groups of the Entisol order) 

Leptosols include soils that are very shallow (less than 25 cm) with continuous rock occurring at or near 

the surface, and soils that are very gravelly (with less than 20 percent soil particles). Bare rock at the surface 

is included in the concept of Leptosols (FAO, 2014; Figure A 5). These are generally young soils with little or no 

soil profile development. When Leptosols form in calcareous materials, dissolution and removal of carbonates 

may occur and biological activity may be high. 

Leptosols may occur everywhere where rocks are near the surface. They are particularly prevalent in strongly 

eroding areas in mountainous land at high and medium altitude with a strongly dissected topography. Minor 

occurrences are also along rivers where gravelly deposits have accumulated without substantial admixture of 

fine earth material. Leptosols are the most extensive soils in the world with an estimated extent of more than 

1 600 million ha. They are associated with mountain ranges, with the Sahara and the Arabian Desert. 

In spite of their considerable extent, Leptosols have largely been ignored in soil studies mainly because 

of their very limited interest for agriculture and the general lack of profile development. This may not be 

fully justified as more than 12 percent of the world’s population lives in a mountainous environment where 

Leptosols are common (Nachtergaele, 2010). 

Leptosols have a potential for seasonal grazing and as forest land. Erosion is the greatest threat in montane 

Leptosol areas of the temperate zone where population pressure (for example from tourism), over-exploitation 

and environmental pollution lead to the increasing deterioration of the natural vegetation and to soil erosion. 


536 


Figure A 5 (a) A Leptosol profile in the 

Northern Ural Mountains and (b) 

associated landscape. 

a 

Photo by Ye. Zhangurov 

b 

Photo by Ye. Zhangurov 


537 


VERTISOLS 

Vertisols (Figure A 6) are expansive clayey soils that shrink and swell extensively with changes in moisture 

content. Cracking, gilgai microrelief, and high clay content are common attributes of Vertisols, but these 

properties are not exclusive to them. Slickensides are the common morphogenetic link to Vertisols and Vertic 

intergrades (Ahmad, 1983; Coulombe et al., 1996a and 2000; Soil Survey Staff, 1999). The soil mechanics model 

of shear failure, slickenside formation, and oblique thrusting appear to better fit observed morphological 

properties, systematic soil property depth functions, and leaching vector transfers than the traditional 

inversion pedoturbation model (Ahmad, 1983; Coulombe et al., 1996a; Wilding and Tessier, 1988). Shrink-swell 

in Vertisols is due mainly to inter- and intra-particle pore volume changes that occur under field conditions at 

high matric potentials (-1/3 to -10 bars). This is in contrast to the commonly held interlayer clay dehydration/ 

rehydration mechanism invoked for shrink/swell dynamics (Wilding and Tessier, 1988). While smectite is a 

clay mineral component commonly found in Vertisols, many other clays including kaolinite, halloysite, mica 

and mixed layer assemblages of vermiculite, smectite and chlorite may occur as dominant or co-dominant 

associates (Coulombe et al., 1996b). Mineralogy controls shrink-swell phenomena by the presence of finegrained 

particles which have high external surface areas, high flexibility and a packing geometry that favours 

micropores a few micrometres in diameter or less (Wilding and Tessier, 1988; Coulombe et al., 1996b). 

Vertisols commonly occur in regions of grasslands and savannas, but may also be found under mixed pine 

and deciduous forests. Parent materials may originate from sedimentary, igneous or metamorphic origins 

but must provide, either from inheritance or weathering, a high content of clay with high surface area. Soil 

moisture conditions vary widely from aridic to aquic with the caveat that soil moisture stress, desiccation 

and cracking must occur at some time in most years. Topography controls gilgai patterns with normal gilgai 

commonly occurring on slopes < 4 percent, while linear gilgai occur on steeper landforms. Most Vertisols occur 

on Pleistocene-age or younger geomorphic surfaces that are several thousand to hundreds of thousands of 

years old (Coulombe et al., 1996a and 2000). 

Vertisols occur in over 100 countries and represent about 316 million km 2 or 2.4 percent of the global icefree 

land area (Dudal and Eswaran, 1988; Ahmad, 1983; Coulombe et al., 1996a; Wilding, 2000; Coulombe et al., 

2000). Over 75 percent of Vertisols are found in India, Australia, Sudan, United States, Chad and China. Fortyseven 

percent are in tropical regions, 52 percent in temperate zones, and 1 percent in cold boreal climates 

(Wilding, 2000; Coulombe et al., 2000). 

While Vertisols are very productive land resources in many parts of the world, especially in the developed 

world, they are among the most difficult resources to manage (Coulombe et al., 1996a; McGarity et al., 1984; 

Ahmad and Mermut, 1996). They require well above average managerial skills for success because of their 

high energy requirements, limited range of soil-water workability, high physical instability, susceptibility 

to seasonal flooding, fertility constraints, and susceptibility to wind and water erosion. They are best 

managed with shallow and infrequent tillage. Irrigation scheduling should be frequent with low application 

rates. Despite their resilience, Vertisols are subject to structural degradation, loss of macroporosity, loss 

of biological diversity and formation of tillage pans when used under continuous, long-term mechanical 

cultivation practices. Rejuvenation of native structural conditions can only be partially achieved after decades 

of fallow (Coulombe et al., 1996a; McGarity et al., 1984; Puentes and Wilding, 1990). Construction activities 

are constrained by the propensity of Vertisols for soil failure and high shrink/swell activity (Coulombe et al., 

2000; McGarity et al., 1984; Ahmad and Mermut, 1996). High bioremediation and physical/chemical sorption 

capacities promote favourable habitats for land treatment of waste products because Vertisols when wet are 

slowly permeable, biochemically reduced, and have long mean residence times for intestinal fluids. 


538 


a 

Photo by L. Wilding 

b 


c 


d 


Figure A 6 Vertisol gilgai patterns and associated soils: (a) linear gilgai pattern located on a moderately sloping hillside in western 

South Dakota. Distance between repeating gilgai cycle is about 4 m. (b) Normal gilgai pattern occurring on a nearly level clayey 

terrace near College Station, TX. After a rainfall event microlows have been partially filled with runoff water from microhighs - 

repeating gilgai cycle about 4 m in linear length. (c) Trench exposure of soils excavated across normal gilgai pattern - repeating 

gilgai cycle about 4 m in linear length. Dark-colored deep soil in microlow (leached A and Bss horizons) with light-colored shallow 

calcareous soils associated with diaper in microhigh (Bssk and Ck horizons). The diaper has been thrust along oblique slickenside 

planes towards soils surface. Vertical depth of soil trench in about 2 m. (d) Close up of dark-colored soil associated with microlow and 

light colored diaper associated with microhigh of the trench in (c). 


539 


SOLONETZ (Natric great groups of several different orders) 

Solonetz (figure A 7 

) are formed by salt accumulation and the leaching of the surface horizon. The dominant 

soil processes involved in Solonetz formation are leaching of the surface horizon combined with argilluviation 

and sodification manifested in the formation of columnar/prismatic structural elements. 

Solonetz generally occur in flat plains that have a source of soluble salts, such as salt-bearing parent 

material or a shallow saline water table in semi-arid, temperate and subtropical climates that receive less 

than 500 mm of rainfall per year. The vegetation is commonly dominated by short grasses. The formation of 

these soils takes generally more than 5 000 years. 

The extreme physical characteristics (high water retention, low hydraulic conductivity, strong swellingshrinking, 

great plasticity) of Solonetz are linked with the high concentration of sodium in the exchange 

complex of the soil. Solonetz show strong profile differentiation in terms of colour, texture, structure, sodicity, 

salinity, alkalinity and calcareousness. Because of the high sodicity, Solonetz have a very short time window 

between snow melt and the following dry period for optimal ploughing. Plastic wet or dry Solonetz surface 

horizons cause a number of problems for cultivation. Only Solonetz with a thick surface horizon can be 

cropped successfully. Other Solonetz may be used for livestock farming. After reclamation, when the adsorbed 

sodium is replaced by calcium (by applying gypsum) and drainage, these soils can be turned into cropland. 

The extent of Solonetz is estimated at about 135 million ha. They are mainly located in North America, 

Eurasia and Australia. 


540 


Figure A 7 (a) A Solonetz profile and (b) 

the associated landscape in Hungary. 

a 

Photo by T. Toth 

b 



541 


SOLONCHAKS (Salids) 

Solonchaks (Figure A 8) are characterized by their high salt concentration, expressed by the electrical 

conductivity of the saturation extract (ECe) that exceeds 15 dS m -1 (or > 8 dS m -1 when the pH is ≥ 8.5). The 

presence of salt crystals and hydromorphic features are indicators of Solonchaks. The dominant soil processes 

involved in Solonchak formation are salt accumulation and the development of hydromorphic features. 

These soils generally occur in inland river basins and very flat or depressed areas which have a source of 

soluble salts, such as salt-bearing parent material or a shallow saline water table. They also occur in coastal 

lowlands. Generally they are formed in arid, semi-arid and sub-humid climates where rainfall is less than 500 

mm yr -1 and the evaporation exceeds the rainfall. The vegetation consists of salt tolerant grassland, bushes or 

mangroves. 

Globally the extent of Solonchaks has been estimated at 260 million ha; they occur mainly in the drier parts 

of North America, northern Africa, the Middle East and central Asia, South America and Australia. 

Solochaks are typically used for livestock farming or highly adapted irrigation farming. The vegetation on 

Solonchaks provides ecological services such as coast protection, grazing land and a source of wood. After 

leaching the salts and with drainage, these soils can be turned into cropland. 


542 


a 


b 


Figure A 8 (a) A Solonchak profile and (b) a salt crust with halophytes. 


543 


4 | Soils distinguished by Fe/Al chemistry 

PODZOLS (SPODOSOLS) 

The majority of Podzols (total area 4.8 million km 2 ) occur in humid boreal and temperate climates on lighttextured 

rocks or quartz sands, on outwash plains and river terraces under pine forests, and on siliceous hard 

rocks in the mountains (Figure A 9). Smaller areas occur under equatorial evergreen forests. 

Podzol is a Russian folk name introduced by Dokuchaev (1879) into scientific language. It means either 

‘similar to ash’ or ‘under ash’, implying that the ash is white or dark, respectively. Hence Podzols initially were 

mostly identified by their whitish (albic) subsurface layer resulting from the loss of iron-organic compounds. 

The thickness of the albic material (and of the whole Podzol solum) depends on climate and ranges from 2 

m in equatorial ‘giant Podzols’ (Sombroek, 1966; Figure A 10) to 5 -10 cm in ‘dwarf Podzols’ on the Baltic Shield. 

Equatorial Podzols are confined to areas with annual rainfall from 1 800 to 3 000 mm without a marked dry 

season, to the weathering products of granites or gneisses, claystone and sandstone in the Amazon basin, and 

to marine sediments on the coastline of Brazil (Sombroek, 1966: Lucas et al., 2012). A particular combination 

of environmental factors favours the development of acid hydrolysis and downward migration of its products 

immobilized at a varying depth to form a spodic horizon. The mechanism of podzolization has been discussed 

by many researchers. The most recent ideas summarized by Sauer et al. (2007). 

The properties of spodic horizons and both the regional and local distribution of Podzols are in good 

agreement with moisture regimes: the spodic horizon is dominated by iron oxide compounds in drier 

conditions and by dark organic matter in humid ones. This differentiation is distinct at regional and local 

levels. Giant Podzols have a high organic carbon content in the spodic horizon, indicating that unusually large 

quantities of dissolved organic carbon were transferred from the topsoil. This is attributed to high volumes 

of water percolating through the soil, the chemical quality of the organic matter, and the long time for soil 

evolution (Lucas et al., 2012). The subsoils of Podzols under a continental climate or strong drainage are 

commonly dominated by iron, while in other more temperate areas or less drained areas they are dominated 

by organic carbon (Friedland et al., 1988). 

Podzols may be young soils, just a few centuries old, or they may have been formed over millennia (Sauer et 

al., 2007). Podzols buried almost 8 000 years ago were described under Histosols 2-3 m thick in West Siberia 

(Karavaeva, 1982). 

Podzols are unstable soils even without human intervention: tree windfalls or fires induce wind erosion. 

Their most efficient ecological services are supporting coniferous forests, often of high quality, regulating the 

water balance in landscapes, and retaining some pollutants. In northern Europe, Podzols on heathlands are 

poorly preserved if they have been part of a Plaggen ecosystem (Blume and Leinweber, 2004). 


544 


Figure A 9 (a) A Podzol profile and (b) an 

associated landscape, West-Siberian 

Plain. 

a 

Photo by D. Kostiuk 

b 

Photo by A. Konstantinov 


545 


Figure A 10 (a) A giant Podzol profile and 

(b) an associated landscape, Brazil. 

a 

Photo by M.R. Coelho 

b 

Photo by M.R. Coelho 


546 


FERRALSOLS (OXISOLS) 

Ferralsols (Figure A 11) form where the weathering conditions are very intense, usually under tropical and 

subtropical humid conditions, with intense leaching of silica and alkaline and alkaline-earth cations, resulting 

in the relative accumulation of kaolinite and various amounts of resistant minerals such as (hydr-)oxides of Fe, 

Al, Mn and Ti. The distribution of clay in the soil profiles is uniform, without marked clay increase with depth. 

Ferralsols have low-activity clays throughout the lower horizons, and the base saturation is frequently low. 

Ferralsols have a distinct granular microstructure, due to the strong interaction among kaolinite and (hydr-) 

oxides of Fe and Al. 

These soils show no large variation of clay content or evidence of clay illuviation, and the horizons are 

marked only by a higher content of organic carbon in the topsoil, which reduces with depth. The subsurface 

horizons show gradual to diffuse boundaries. The chemical characteristics reflect the leaching of base cations 

and advanced weathering resulting in low-activity clays. Some Ferralsols formed from basic rocks such as 

basalt may have better nutrient reserves, although the high iron content will result in strong phosphorus 

‘fixation’. The structure is usually of granular type, although some Ferralsols may develop a weak subangular 

block structure. They vary in the colour of their subsurface horizon from red to yellow, mainly according to the 

iron content in the parent material and to hydrological conditions. 

Ferralsols are reported on weathering products of acid and basic rocks, and unconsolidated sediments on 

old and stable surfaces. They are most common on interior plateaus or slowly undulating topography in humid 

tropical, humid subtropical and monsoon climates. Because many climatic changes occurred since these soils 

were formed or the parent materials deposited, they may lack a relationship with the present vegetation, 

which may vary from Amazon forest to dry savannah. 

The most extensive occurrences of Ferralsols are in South America, mainly in Brazil. They cover about 17 

percent of Latin America and the Caribbean (Gardi et al., 2014). Ferralsols are also distributed in eastern and 

central Africa (10 percent of the continent, Jones et al., 2013) and Madagascar, in some areas of Australia and 

in the United States (Hawaii). 


547 


a 

Photo by M.G. Pereira 

b 

Photo by F.H. Gomez 

Figure A 11 (a) A Ferralsol profile and (b) an associated landscape, Brazil. 


548 


NITISOLS (Alfisols, Ultisols, Inceptisols and Oxisols Great groups) 

Nitisols are well drained clayey soils with deep profiles. They are characterized by the strong development of 

structure, frequently with shiny aggregate faces (Figure A 12). They originate from basic and intermediate rocks 

or sediments derived under relatively intense weathering conditions in tropical and subtropical climates. This 

leads to the predominance of low activity clays (kaolinite) and (hydr-) oxides of Fe, Al and titanium (Ti). Some 

Nitisols have high base saturation and high potential for crop production. Others have very high amounts of 

iron and strong P fixation, or a very low sum of exchangeable bases and high aluminium. Both these latter 

classes are limiting to crops. 

The texture is clay loam or finer, with no large variation of clay content within the soil. The profile development 

shows intense weathering, with the prevalence of kaolinite and high iron in the nitic horizon, resulting in the 

strong stability of the aggregates, and the common angular and/or subangular blocks combined in a prismatic 

structure. The nitic horizon may show clay coatings indicating an illuviation process. The Nitisols are usually 

red or reddish-brown, and there is no distinct colour variation in the profile, except for the topsoil, due to the 

higher content of organic carbon. The subsurface horizons show gradual to diffuse boundaries. Nitisol classes 

vary largely according to base saturation, clay activity (usually low), iron and aluminium content. However, 

Nitisols formed from basalt may have high base saturation and, due to their good drainage and structure, may 

have high potential for both intense and low input agriculture. 

Nitisols are mainly formed from weathering products of intermediate and basic igneous rocks (basalt 

and diabase). They may also have originated from clayey sediments in karstic areas (TerraRossa). They occur 

predominantly on high level plateaus and slightly undulating reliefs, originally under tropical and subtropical 

forest, or Cerrado (Brazil) and savannah vegetation. 

Nitisols occur in eastern Africa and Madagascar (2 percent of the continent, Jones et al., 2013). Although 

accounting for less than 1 percent of area on the Latin America and Caribbean soil map, Nitisols are prized 

lands in Southeastern and South regions of Brazil, and in neighbouring Argentina and Uruguay. They are 

cultivated with crops such as coffee, citrus, soybean, corn and sugarcane; and they play an important role 

in the agriculture of many tropical countries. Nitisols are also found in Australia (Ferrosol in Australian soil 

classification, formed from basalt), Europe (the Mediterranean) and the United States. 


549 


a 

Photo by L. Anjos 

b 

Photo by L. Anjos 

Figure A 12 (a) A Nitisol profile and (b) the associated landscape with termite mounds, Brazil. 


550 


PLINTHOSOLS (Plinthic sub-groups) 

Plinthosols (Figure A 13) are defined by the presence of plinthic, petroplinthic or pisoplinthic horizons, 

at a certain depth in the soil profile. Their formation is related to accumulation and redistribution of Fe 

under conditions of alternating wetting and drying cycles over long time periods. The landscape position 

- low lands with high groundwater or slopes with water seepage conditions - leads to chemical reduction 

of iron compounds in the parent material, which are redistributed and accumulated in the soil profile. The 

plinthite may be hard and irreversible (petroplinthite), forming a continuous and highly impermeable layer 

of ferruginous material (carapace or crust). The reduction, segregation and precipitation of iron (hydr-)oxides 

in the subsurface horizon forming the plinthite bodies, together with the dominance of kaolinite and other 

products of strong weathering such as gibbsite, indicate the conditions in which most Plinthosols formed. 

The profile may develop strongly bleached eluvial horizons and have evidence of clay illuviation; or show 

morphology associated to Ferralsols or to lesser development in recent sediments where reducing conditions 

are still present as indicated by gleyic properties. The subsurface horizon has platy, polygonal or reticulate 

patterns of distinct coloured (red, brown) plinthite bodies that are coherent enough to be separated from 

the surrounding soil matrix, which is usually of a pale colour. Hardening of the plinthite will form discrete 

concretions or nodules that characterize the pisoplinthic horizon. Further cementation and interconnecting 

of the pisoplinthic material will form the petroplinthic horizon, a layer of indurated material which may be 

continuous, broken or fractured. 

Plinthosols are reported as formed from weathering products that have a high amount of Fe or where this 

element is accumulated due to water seepage or ascension of groundwater. They are most common on level 

to gently sloping topography, in areas with seasonal fluctuating groundwater in wet climates, humid and 

tropical, such as in the Brazilian Amazon Basin. However, in the Brazilian Cerrado and the savannahs of Africa, 

Plinthosols (with petroplinthic or pisoplinthic horizons) are also found on steeper slopes or as hard layers on 

plateau tops of old erosional surfaces. 

Extensive areas of Plinthosols occur in West Africa, where they represent 5 percent of the total 30 million 

km 2 area of the continent (Jones et al., 2013). Widespread in the Amazon Basin, they cover about 1 percent of 

Latin America’s 22 million km 2 , largely in Brazil, Colombia, Venezuela, Guyana and Bolivia, and in the Caribbean 

region, (Gardi et al., 2014). Plinthosols are also found in Southeast Asia, India, Australia and the United States. 


551 


Figure A 13 (a) A Plinthosol profile, (b) 

details of the plinthic horizon and (c) the 

associated landscape, South Africa. 

Photo by ISRIC World Soil Information. 

a 

b 


c 



552 


PLANOSOLS (Albaqualfs, Albaquults and Argialbolls) 

Planosols are seasonally water-saturated or flooded, poor acid soils with bleached, generally silty surface 

horizons with an abrupt transition to a dense subsoil with significantly more clay (Figure A 14). There may be 

pore infillings of bleached material in the subsoil. Clay destruction and aluminium interlayering driven by 

periodic iron hydroxide reduction and reoxidation (ferrolysis) has been recognised as a process sometimes 

involved in the formation of the silty surface horizons (Brinkman, 1979; Van Ranst et al., 2011). 

Planosols occur in generally level areas in climates with contrasting wet and dry seasons, mainly in the 

subtropics but in temperate areas and the tropics as well. Their total extent is estimated at 1.3 million km 2 . 

They are extensive in Latin America (southern Brazil, Paraguay, and Argentina) and Australia, and they also 

occur in Africa (Sahelian zone, East and southern Africa), the eastern United States, Siberia, China, and 

Southeast Asia (Bangladesh, Thailand). 

Natural vegetation on Planosols is sparse grass with or without shrubs or small trees; extreme Planosols 

may be barren. They are generally used for grazing or for grain or root crops in temperate areas. In the 

subtropics and tropics, rainfed paddy (wetland) rice is grown on bunded fields; with irrigation, they can be 

double cropped with a second paddy rice or dryland crop. Yields are very low without fertilizers and remain 

sub-optimal even with fertilizers because of the poor physical and chemical soil conditions. 


553 


a 

Photo by National Institute of Agriculture Technology (INTA), Argentina 

b 

Photo by National Institute of Agriculture Technology (INTA), Argentina 

Figure A 14 (a) A Planosol profile and (b) the associated landscape, Argentina. 


554 


GLEYSOLS (Aquic suborder and Endoaquic great groups) 

Gleysols are easily identified by bluish or greenish grey colours in their mineral horizons that are usually 

water-saturated, with only a weak or no structure (Figure A 15). These horizons are formed under reducing 

conditions characterized by a low redox potential. In some Gleysols the smell of hydrogen sulphide or methane 

is noticeable. Iron compounds are easily mobilized in Gleysols, especially in the presence of organic matter and 

anaerobic microorganisms. These are partially removed or oxidized and may accumulate as iron segregations, 

nodules, iron pans, bog ores, etc. (Zaidelman, 1994). Above the layer with gleyic characteristics, a topsoil 

horizon relatively rich in organic matter occurs, that may show rusty root channels. The range in pH values in 

Gleysols is broad and may vary between 2.5 and 9. In coastal positions Gleysols may show sulphides oxidation 

resulting in high acidity (Zech et al., 2014). 

Globally the extent of Gleysols is estimated at 7.2 million km 2 , of which approximately two thirds occur in 

boreal areas on unconsolidated parent rocks. In humid regions they often occupy depressions, river valleys and 

deltas, lake kettles and foot slopes. Subaqueous soils of shallow water bodies are also included with Gleysols. 

Large areas of Gleysols occur in tundra areas, in deltas of great rivers and in lowlands. They occur as associated 

soils almost everywhere, except in arid lands and on steep slopes. 

In tundra regions the melting of the permafrost layer in summer causes excess of water in an environment 

already enriched in organic matter and induces seasonally reducing conditions and the formation of Gleysols. 

Water logging is the main prerequisite for the development of gleyic features and is due to high ground water 

table in depressions; additional water inflow there may contribute to gleying as well as flooding in the valleys 

and tides in coastal areas. There is no special plant community on Gleysols because they occur all over the 

world, but everywhere hygrophytes are dominant plants. 

The main limitation for Gleysols management is surface water logging and/or shallow ground water 

hindering the growth of the roots of crops and trees. With artificial drainage the ground water table is lowered 

and the excessive moisture removed. When drainage is implemented efficiently, as it is in the Netherlands and 

Germany, Gleysols are productive soils for vegetables, beets and flowers. The main ecosystem threat is related 

to the Gleysols’ low position in the landscape, where they may accumulate pollutants and could turn into 

‘chemical time bombs’ (Stigliani, 1988). 


555 


a 


b 


Figure A 15 (a) A Gleysol profile and (b) associated landscape in the East European tundra. 


556 


STAGNOSOLS 

(Aquic Suborders and Epiaquic Great Groups in Alfisols, Ultisols, Inceptisols, Entisols, and Mollisols) 

Stagnosols (Figure A 16) have much in common with Gleysols, but are different in their source of waterlogging 

and in their manifestations in the soil profile. Periodical stagnation of atmospheric water accounts for the 

name of this soil (originating from the German Pseudogley and Stagnogley). Stagnosols are characterized by 

the difference in texture between topsoil and subsoil originated due either to illuviation or to initial parent 

material heterogeneity in areas with humid climate and flat topography. Stagnosols are identified by the 

colour pattern of the upper 0.5 m of their mineral horizon, where a combination of reductimorphic (bluish 

grey colours that do not last) and oximorphic colours (rusty, reddish brown mottles inside aggregates and 

root channels known as Rohrenstein) together with iron-manganic segregations or nodules occur. These 

pedofeatures may occur within the whole layer, or they may be confined to its lower part, whereas its upper 

part may be composed of albic material with reductimorphic features. A special case of stagnic properties is 

the ‘marbled’ colour pattern described in old German literature as ‘Marmorierung’ (Muckenhausen, 1963). 

Stagnosols are mostly acid to weakly acid and have a low to medium base saturation. Humus accumulation 

is prominent in these soils with raw or moder humus types; the biological activity in these soils is weak and 

the physical properties are unfavourable for plant growth: low porosity, reduced water filtration and risks 

of drying out (Zech et al., 2014). Stagnosols are often localized and do not occur in vast continuous areas. 

They are mostly associated with other soils - Cambisols, Retisols, Acrisols. They are confined to flat or weakly 

undulating plains with various unconsolidated parent materials, moderately or heavy-textured. When the 

textural difference between the top- and subsoil is large, they are replaced by Planosols. 

Stagnosols have mostly been described in areas with humid temperate and subtropical climate 

under hardwood forests. They are most common in Western Europe and the Midwest of the 

United States. The total area of Stagnosols worldwide is estimated at 1.5 -2 million km 2 (FAO, 2014). 

Stagnosols have a low fertility due to their poor physical properties and moisture regime along with the 

elevated acidity and aluminium toxicity. Applying artificial drainage is less efficient than in Gleysols, unless 

additionally deep loosening of the subsoil is applied to break the impermeable layer. The same weakly 

permeable and dense subsoil is a problem for silviculture as it is an obstacle for tree roots and results in a 

high probability of tree uprooting. Nevertheless, forests of wetness-tolerant tree species and meadows are a 

preferable land use option. 


557 


Figure A 16 (a) A Stagnosol profile, (b) 

stagnic color patterns, (c) marble-like 

horizontal surface and (d) an associated 

landscape. 

Photo by M. Gerasimova 

a 

b 


d 


c 



558 


ANDOSOLS (ANDISOLS) 

Andosols (Figure A 17) form typically from volcanic tephra on uplands. Distinctive properties of matured 

Andosols are the high content of active Al and Fe materials, and the lowest bulk density among mineral 

soils. Volcanic glass, the major constituent of tephra, rapidly weathers to form allphane, imogolite, Alhumus 

complex (non-crystalline active Al materials) and ferrihydrite (poorly crystalline active Fe material), 

with leaching loss of a large amount of Si, Na, Ca, etc. Under plentiful vegetation a large amount of humus 

accumulates in the A horizon, forming the Al-humus complex. Halloysite tends to increase under semi-dry 

climate in addition to non-crystalline active Al materials. Translocation of clays, Al and Fe is minimal in the soil 

profile. Due to a porous, fluffy and highly aggregated microstructure, Andosols show low solid phase ratio, 

low bulk density, high water permeability, and high water holding capacity. 

A humid climate and uplands are favourable for leaching loss of Si, Na, Ca, etc. and for formation of active 

Al and Fe materials. Tephra deposits in swamps tend to weather more slowly than those on uplands, and the 

weathering product is richer in halloysite. Weathering of tephra under arid soil moisture regime appears even 

slower. The rock type of tephra ranges from rhyolitic to basaltic. The colour of rhyolitic to Andisitic tephra is 

whitish to greyish and that of basaltic tephra, black. 

Andosols cover less than 1 percent of the earth’s land surface. The major occurrences of Andosols are in 

and around the volcanic areas along the circum-Pacific volcanic zone, the Alpine-Himalayan belt, and the 

great Rift Valleys of Africa. Others are on the Hawaiian Islands, Iceland, etc. Andosols are used as productive 

farmlands after appropriate improvement of chemical shortcomings. 


559 


Figure A 17 (a) An Andosol profile and (b) 

the associated landscape in Japan. 

a 

Photos by M. Nanzyo 

b 

Photos by M. Nanzyo 


560 


5 | Soils with accumulation of organic matter in the topsoil 

CHERNOZEMS (Udolls) 

Chernozems are marked by their deep, dark and well-structured topsoil (FAO, 2014). The soil has an almost 

black colour (hue < 1, chroma ≤ 2.5), intricate pedality and strong water-stable structure, which is mostly 

due to the activity of earthworms. The (micro-)structure is granular or crumb-granular in the upper part of 

the topsoil and spongy in the lower one; density is close to 1 g cm -3 ; the Corg content ranges between 2.9 

and 3.5 percent in the upper 10 cm (with humates as predominant fraction), and exceeds 1.2 percent at 

the lower boundary of the chernic horizon (Lebedeva, 1974). Earthworms and burrowing mammals (mole 

rats, marmots, hamsters, ground squirrels) modify the horizons’ boundaries, effervescence depth and the 

pathways of solution flows. They also perform the exchange of material between the top and subsoil, which 

contributes to the profile stability; dark and brown krotovinas are common. Calcic horizon and/or secondary 

carbonates are diagnostic for all Chernozems. Secondary carbonates comprise labile forms: pseudomycelium 

and impregnation mottles corresponding in thin sections to needle-shaped crystals in voids, micritic (quasi) 

coatings, sometimes with sparite grains. The labile forms of carbonates are in agreement with the data on 

current hydrothermal soil regimes. Soft segregations – beloglazka – and hard nodules occur in more arid 

variants of Chernozems transitional to Kastanozems, while micritic pendants are confined to materials 

with rock fragments. In Russia, Chernozems are differentiated in accordance with secondary carbonate 

pedofeatures reflecting the current pedoclimate (Figure A 18). 

Continental climate with summer rains, soil freezing for two to four months, rich forb-grass natural 

vegetation, mostly loess as parent material, good drainage and level to undulating topography all contribute 

to the development of most typical profiles. The radiocarbon age of the topsoil ranges within 2 -3 kA in its upper 

part, and 5-8 kA in the lower part (Chichagova, 1985). Chernozems first appeared in the Late Miocene under 

grass ecosystems maintained by grazers (Retallack, 2001). However, most Chernozems have been cropped 

for at least the last two centuries. They are regarded as very fertile soils. 

Chernozems occur as a continuous belt in steppe and forest-steppe landscapes in Russia and the Ukraine, 

in the Great Plains of the US, in northern Kazakhstan and locally in some countries of Central Europe. They 

cover approximately 230 million ha. 

High fertility of Chernozems is provided by a unique combination of very favourable chemical and physical 

properties. More than half of their area is cropland - maize, wheat, sugar beet and sunflower are the main 

crops. In the drier parts of their area, the main limitations to agriculture are droughts with occasional dust 

storms, whereas in wetter parts both wind and water erosion are the main risks. Climate change along with 

water conservation measures and irrigation at the background of lithological discontinuity have resulted in 

the appearance of small wetlands in the steppe landscapes. 


561 


Figure A 18 (a) A Chernozem profile (Photo 

by J. Deckers) and (b) the associated 

landscape in the Central Russian 

Uplands. 

a 


b 



562 


KASTANOZEMS (Ustolls and Xerolls) 

Kastanozems are humus-rich soils that were originally covered with early-maturing native grassland 

vegetation which produced a characteristic brown topsoil 20-40 cm thick in which the organic matter 

content ranges between 2 and 6 percent. Kastanozems have a brown topsoil with a granular or fine blocky 

structure. The rest of the profile is lighter in colour and is characterized by the secondary accumulation of 

calcite (Figure A 19). Kastanozems are chemically rich soils with a pH slightly above neutral. Near the surface, 

soil pH may reach a value of 8.0. 

These soils are found in relatively dry climatic zones (annual precipitation 200–400 mm). Kastanozems 

are mostly used for irrigated farming and grazing. Kastanozems have relatively high levels of available calcium 

ions and other nutrients. Carbonates weakly move down in the soil profile with percolating water to form 

layers of secondary carbonates; gypsum is also common in these soils. Kastanozems form in semi-arid regions 

under relatively sparse grasses and shrubs. 

The total extent of Kastanozems is estimated to be about 465 million ha. Major areas are in the Eurasian 

short-grass steppe belt (southern Ukraine, the south of the Russian Federation, Kazakhstan and Mongolia), 

in the Great Plains of the United States of America, in Mexico, and in the southwestern pampas and Chaco 

regions of Argentina, in Paraguay and southeastern Bolivia (FAO, 2014). 

The main obstacle to the agricultural use of these potentially rich soils is drought (Encyclopaedia of Soil 

Science, 2008). Irrigation, which brings the threat of secondary salinization, is nearly always necessary to 

obtain high yields. Another serious problem on Kastanozems is overgrazing (Wang and Batkhishig, 2014), 

extensive grazing being another important use for these soils. Overgrazing on light-textured soils often 

produces deflation, destroying the topsoil. 


563 


Figure A 19 (a) A Kastanozem profile and 

(b) the associated landscape in Mongolia. 

a 

Photo by O. Bathkhishig 

b 

Photo by O. Bathkhishig 


564 


PHAEOZEMS (Udolls and Albolls) 

Phaeozems are soils with a mollic horizon which occur most frequently in the transitional areas between 

boreal forests and steppes, or in forest-free plains with temperate semi-humid climate (tall-grass prairies) on 

unconsolidated base-rich sediments, mostly loess or loess-like material. They may also occur locally under 

sparse herbaceous forests in the mountains (Figure A 20). They cover approximately 2 million km 2 , and their 

largest areas are found in the United States (Great Plains) and Canada, in the Argentinian Pampas, and in 

Manchuria. Mountainous variants were described in Southern Siberia and Northern Mongolia on gentle 

slopes with colluvium, under larch forests with a rich forb-grass cover (Zech et al., 2014; Vostokova and Gunin, 

2005). Phaeozems are formed under milder and more humid climates than Chernozems; the vegetation is 

mesophytic with less pronounced seasonal rhythms. Typically, natural grassland cover is mostly replaced by 

high-quality farmland, or may be modified by grazing. Phaeozems are among the most fertile soils owing to 

their favourable physical and chemical properties, along with moisture and thermic regimes (udic and mesic). 

The most conspicuous feature of Phaeozems is their dark, mostly thick, mollic horizon with traces 

of burrowing mammal activity, weakly acid to neutral, base saturation ranging within 50-100 percent. 

Phaeozems include some of the traditional forest-steppe Chernozems with deep secondary carbonates, and 

in this case they have a chernic horizon with its coprogenic structure underlain by a cambic or argic horizon. 

In the rest of Phaeozems, the subsoil horizons may be diverse: argic, cambic, calcic and petrocalcic; the latter 

phenomenon is common in Argentinian Phaeozems – a specific hard tosca layer that may occur within 1 m 

from the soil surface and be a limitation for plant growth (Moscatelli, 1991; Pazos, 2012). Some other properties 

were described in Phaeozems as well: albic material and uncoated silt grains, clay coatings, stagnic colour 

pattern, and sodic features (FAO, 2014). This broad array of properties is explained by the occurrence of 

Phaeozems in different environments providing for the development of additional pedogenic processes, some 

of them being limitations for farming. 

The global significance of Phaeozems is their high agricultural potential, as well as the prominent reserves 

of organic carbon accumulated in their topsoils. The limitations are not strong: they include wind erosion in 

dry years, water erosion on uplands, and water stagnation either during short rainy events or in case of high 

groundwater. In Manchuria, deep freezing and slow thawing are common. Local manifestations of sodicity 

have been recorded in Argentina and Western Siberia (Gerasimova, 2002). 


565 


Figure A 20 (a) A Phaeozem profile and (b) 

the associated landscape, Argentinian 

Pampa. 

a 

Photo by S.M. Pazos 

b 

Photo by S.M. Pazos 


566 


UMBRISOLS (Umbric Great Group in Aquept Suborder, Humic Subgroups in all Suborders of Inceptisols) 

Umbrisols are mostly mountainous soils of cool humid climates covered by meadows or sparse forests. 

They are characterized by a dark, humus-rich and acid topsoil horizon with a low base saturation. A rather 

weak crumb structure is characteristic for the topsoil horizon (Figure A 21). 

Umbrisols are formed under dense forb-grass natural vegetation (subalpine meadows) or under deciduous 

forests with a prominent lower canopy, sometimes with shrubs. This produces a large volume of plant residues, 

which in part may not be strongly decomposed, and elements of a moder humus form may be identified (Zech 

et al., 2014). Rather steep slopes and stony parent material provide sufficient drainage in spite of abundant 

precipitation and high air moisture; the soil is always moist, but stagnic or gleyic properties are absent. Typical 

examples of landscapes with Umbrisols are (sub-)tropical montane cloud forests in Mexico, Bolivia and Chile 

(Roman et al., 2010), although Umbrisols also occur at higher altitudes in sub-boreal continental mountain 

ranges. Igneous and metamorphic rocks are almost always the parent material for Umbrisols. Worldwide, 

Umbrisols occupy approximately 10 million km 2 (Zech et al., 2014). 

The geographical location of Umbrisols poses serious limitations for agricultural activities. Chemical 

fertility is not low owing to high humus content but is restricted by soil acidity. Liming and mineral fertilizers 

are required. Another limiting factor is the risk of erosion because of the predominance of steeper slopes in 

areas of Umbrisols. Most Umbrisols are left under natural forests or forestation activity as hard rock or stony 

eluvium are not serious obstacles for tree roots. Grazing is less common. Only in New Zealand have high inputs 

made it possible to practice intensive dairy farming on these soils (FAO, 2014). 


567 


Figure A 21 (a) An Umbrisol profile, 

(b) associated vegetation and (c) an 

associated landscape. 


a 

c 


b 



568 


6 | Soils with accumulation of moderately soluble salts 

DURISOLS (Durids) 

Durisols may develop in arid and semi-arid conditions when a dissolution and accumulation of silica leads 

to the formation of a cemented hardpan that restricts the rooting depth of soils. These soils form mainly in 

alluvial and colluvial deposits in level or slightly sloping alluvial plains, terraces and piedmont plains. Stable 

landscapes occur where the Durisols have been eroded down to their resistant duripan, the material of which 

is often used in road construction. Durisols in low-lying areas may suffer from salt accumulation (Figure A 22). 

The duripan may range in thickness from 10 cm to more than 4 metres. There are two main types of duripans: 

those which are massive, and those with a platy or laminated structure that are coated with amorphous opal 

or microcrystalline silica. 

Durisols are known to be relatively extensive in Australia, South Africa, Namibia and the drier parts of the 

southern United States. Minor extents have been observed in South America and Kuwait. No estimate of their 

global extent is available (FAO, 2014). The agricultural use of Durisols is mostly limited to extensive grazing. 

Arable cropping is limited to areas where irrigation water is available. 


569 


Figure A 22 (a) A Durisol profile and (b) the 

associated landscape, Ecuador. 

a 


b 



570 


CALCISOLS (Calcids, Argids, Cambids, some Cryids) 

Calcisols embrace a broad group of soils in arid and semiarid regions. Their name in the FAO-WRB system 

has been changed, with former Xerosols and sub-categories of Yermosols becoming Calcisols. Most national 

names for these soils comprise indications of their grey or brown colour and of the (semi-)desertic or aridic 

landscapes in which they occur (Figure A 23). Calcisols are widely spread in Mediterranean countries, centraleastern 

and southern Africa, the Near East, Mongolia, Australia, and southwestern United States. Calcisols 

are very widespread and are fifth in importance by surface area of all classified soils - 1000 million ha (FAO, 

2014). 

Calcisols have light coloured topsoil, poor in humus, sometimes free of carbonates, and a diagnostic calcic 

horizon. If there is a petrocalcic horizon within the upper 100 cm, the soil is also qualified for Calcisol. Calcic 

horizon is identified in the profile either ‘quantitatively’ by an elevated content of calcium carbonate in the 

fine earth ( ≥ 15 percent CaCO 3 

), or by an increase relative to the underlying horizon; or ‘qualitatively’ through 

the presence of secondary carbonates (FAO, 2014). Both criteria indicate mobilization and accumulation of 

carbonates in the soil (calcification). Both processes are known to depend on the moisture regime (Boettinger, 

2002), which is dry almost all year round but with a short rainy period. Calcic horizon is formed in other soils 

(salt-affected, Chernozems, Kastanozems) but in Calcisols it is their major characteristic. High content of 

carbonates may be checked in the field by effervescence with 1M hydrochloric acid: it is quick with abundant 

foam formed. Secondary carbonates occur as soft nodules (beloglazka), pendants and coatings on stones, 

impregnation mottles, veins, single or coalescent – pseudomycelium, in the fine earth. Calcisols are always 

base-saturated, neutral to alkaline, have a narrow C:N ratio, and Corg content below 1-2 percent; the profile 

curve of CaCO 3 

usually has a peak in the subsoil (Zech et al., 2014). 

Water deficit is a major limitation for using Calcisols, and extensive grazing is common in many lands 

dominated by Calcisols. Few areas are used for rainfed agriculture. Under irrigation, grain crops, cotton and 

vegetables are efficiently grown. 


571 


Figure A 23 (a) A Calcisol profile, (b) an 

associated landscape and (c and d) 

secondary carbonates in Calcisols. 


a 


c 

d 


b 



572 


GYPSISOLS (Gypsids) 

Gypsisols are characterized by a significant secondary accumulation of calcium sulphate. Accumulation 

of gypsum takes place initially as crystal aggregates in the voids of the soils. These aggregates grow 

by accretion, displacing the enclosing soil material. When the gypsic horizon occurs as a cemented 

impermeable layer, it is recognized as the petrogypsic horizon. These soils occur in the driest part of the arid 

climatic zone in unconsolidated deposits of base-rich weathering material on level land and in depressions. 

Natural vegetation on these soils is sparse and limited to xerophytes and ephemeral grasses and herbs 

(Figure A 24). 

The worldwide extent of Gypsisols has been estimated at about 100 million ha, exclusively occurring in 

desert areas. Major occurrences are found in the Near East, Kazakhstan, Turkmenistan, Uzbekistan, the 

Libyan and Namib deserts, in southern and central Australia and in the southwest of the United States. 

Large areas of Gypsisols are used for extensive grazing. When irrigation water is available these soils can 

be very productive, but the dissolution of gypsum results in the irregular subsidence of the land surface, 

caving in canal walls, and in the corrosion of concrete structures (FAO, 2014). 


573 


a 

b 

Photos by M. Gerasimova Photos by M. Gerasimova 

Figure A 24 (a) A Gypsisol profile and (b) an associated landscape. 


574 


7 | Soils with a clay-enriched subsoil 

RETISOLS (Glossic great groups of Alfisols and Ultisols) 

The clay illuviation within Retisols is typically manifested by an interfingering of bleached coarser-textured 

soil material into the illuvial horizon, forming a net-like pattern (e.g. a glossic horizon). The dominant soil 

processes involved in Retisol formation are argilluviation and biological enrichment of base cations. They are 

often characterized by ‘waxy’ argillans; a subangular blocky structure; silty or loamy textural classes; active 

and superactive CEC (cation-exchange capacity) classes; and the occurrence of lithologic discontinuities 


Retisols occur in climates where winters are cold and summers are short and cool with an annual 

precipitation between 500 and 1000 mm. They typically carry a temperate needle-leaf evergreen forest/ 

woodland on often steeply sloping land. Their parent material is variable and includes loess, till, lacustrine 

and alluvium. Retisols are dated from the mid-Holocene or older e.g. > 5 000 years old. These soils generally 

exist in ‘tension zones’ (ecotones), reflecting a change in climate and/or vegetation. 

Regional distribution of the 320 million ha of Retisols is mainly in Europe and northern and central 

Asia. There are about 85 000 ha in the United States. Retisols are important for forestry, recreation, and 

limited livestock farming and they provide ecological services such as watershed protection and ecological 

sustainability 


575 


Figure A 25 (a) A Retisol profile, (b) the 

“retic” pattern in a Retisol and (c) the 

associated landscape, Belgium. 

Photo by C. Marsboom 

a 

b 


c 



576 


ACRISOLS (Kan- great groups of Ultisols, e.g. with a kandic horizon) 

Acrisols are characterized by movement and accumulation of low-activity clays (cation-exchange capacity 

< 24 cmolc kg -1 clay) and a low base saturation (< 50 percent). The dominant soil processes involved in 

Acrisol formation include argilluviation and base-cation leaching (Figure A 26). 

Acrisols occur under equatorial or warm climates, fully humid or winter-dry with an annual precipitation 

exceeding 1 200 mm. They typically carry a tropical deciduous or tropical evergreen forest or are under 

savannah. They occur on old hilly land surfaces where the relief is variable but often steeply sloping. Their 

parent material is saprolite or colluviums. These soils are commonly more than 200 000 years old. 

Acrisols are used for forestry, recreation, agroforestry and shifting cultivation. They provide ecosystem 

services such as water protection and biotechnology for human health. 

Regional distribution is some 1 000 million ha worldwide, mainly in southeast Asia, the southern fringe 

of the Amazon Basin, southeastern United States, and east and west Africa. 


577 


a 


b 


Figure A 26 (a) An Acrisol profile and (b) the associated landform in Kalimantan, Indonesia. 


578 


LIXISOLS (Kan - great groups of Alfisols, e.g. with a kandic horizon) 

Lixisols are characterized by the movement and accumulation of low-activity clays (cation-exchange 

capacity < 24 cmolc kg -1 clay) and a high base saturation (> 50 percent). The dominant soil processes 

involved in Lixisol formation include argilluviation and biological enrichment of base cations. These soils are 

often polygenetic and have strong textural differentiation and advanced weathering but with abundant base 

cycling (Figure A 27). 

Lixisols occur in the drier parts of the tropics and sub-tropics with a precipitation more than 1 200 mm 

annually. They typically carry a savannah vegetation. They occur on variable reliefs, while their parent 

material is saprolite or colluviums. These soils are commonly more than 200 000 years old. 

Regional distribution is 435 million ha worldwide, mainly in sub-Sahelian and east Africa, Central and 

South America, the Indian Subcontinent, and southeast Asia and Australia 

Lixisols are used for forestry, low-volume grazing and agro-forestry and provide ecological services such as 

water protection and ecological sustainability. 


579 


Figure A 27 (a) A Lixisol profile and (b) the 

associated landscape, Brazil. 

a 


b 



580 


ALISOLS (Ultisols with an argillic horizon) 

Alisols are characterized by movement and accumulation of high-activity clays (cation-exchange capacity 

> 24 cmolc kg -1 clay) and a low base saturation (< 50 percent). The dominant soil processes involved in 

Alisol formation include argilluviation and base-cation leaching. Alisols have high Al and very low plant 

nutrients (Figure A 28). 

Alisols occur under equatorial or warm climates, fully humid or winter-dry with an annual precipitation 

exceeding 1 200 mm. They typically carry a tropical deciduous forest or tropical evergreen forest. They 

occur where the topography is variable but often hilly or undulating, while their parent material is strongly 

weathered basic rocks and unconsolidated sediments. These soils are commonly more than 200 000 years 

old. 

Regional distribution of the approximately 100 million ha of Alisols globally is mainly in Central and 

South America, the Caribbean, west and east Africa, southeastern Asia, and northern Australia. Alisols are 

used for forestry, low-volume grazing and, to a limited extent, for agriculture. 


581 


Figure A 28 (a) An Alisol profile and (b) the 

associated landscape, Belgium. 


a 

b 



582 


LUVISOLS (Alfisols with an argillic horizon) 

Luvisols are characterized by clay movement, accumulation of high-activity clays (CEC > 24 cmolc kg -1 

clay) and a high base saturation (> 50 percent). The dominant soil processes involved in Luvisol formation 

include argilluviation and biological enrichment of base cations. They are either derived from base-rich 

materials or have not been subject to strong weathering (Figure A 29). 

Luvisols occur in humid climates with warm summers and snowfall during winter. They typically carry a 

vegetation of deciduous forest or woodland. They occur on flat or gently sloping topography. Their parent 

material is till, loess, alluvium or colluvium. These soils are commonly more than 5 000 years old. 

There are 500-600 million ha of Luvisols worldwide, mainly in the Eastern European Plain, Western 

Siberian Plain, north central and northeastern United States, central Europe, and South Australia. Luvisols 

are used for agriculture, forestry and grazing. They are among the most productive soils worldwide and 

provide ecological services such as food and energy security, water protection, and ecological sustainability. 


583 


Figure A 29 (a) A Luvisol profile and (b) the 

associated landscape, China. 

a 


b 



584 


8 | Soils with little or no profile development 

These soils have little or no profile development due to age, parent material, soil depth, transport, or 

deposition. 

CAMBISOLS (Inceptisols) 

Cambisols are young soils with beginning subsurface soil development. Characteristics that are more 

easily modified include structure, colour and bulk density. Structure begins to develop as wetting and drying 

cycles occur. Colour is modified through additions and removals such as carbonates and silica. Bulk density 

decreases as elements are weathered and organisms create voids. Typical soil horizonation is A-Bw-C 


Cambisols are found in a wide range of climates, in all vegetation types, and level to steep reliefs. The 

typical parent material is medium and fine-textured, derived from a wide range of rocks, mostly in colluvial, 

alluvial or aeolian deposits. Cambisols form in almost all environments except permafrost. 

The spatial distribution of Cambisols is estimated to be 1 500 million ha worldwide. Countries with 

more than 50 million ha are Russia, China, Canada and India. Cambisols are the dominant soil in San 

Marino, Saint Lucia, Grenada, Saint Vincent and the Grenadines, Jersey, Fiji, Belize, Italy, Luxembourg, 

Samoa, Guernsey, Anguilla, Czech Republic, Georgia, Haiti, American Samoa, Solomon Islands, Bosnia and 

Herzegovina, and New Zealand. 


585 


Figure A 30 (a) A Cambisol profile and (b) 

the associated landscape, China. 

a 


b 



586 


REGOSOLS (Orthents) 

Regisols are the youngest soils with no pedogenic horizons and no evidence of soil forming processes. 

They may, nonetheless, support vegetation but do not meet criteria to be classified as another soil. They are 

located in inert or slowly soluble parent material, recent deposits, or excavation spoils. The profile horizons 

are usually A-C (Figure A 31). 

Regisols exist in all climates and vegetation. They occur on level terrain to steep slopes. 

Countries with more than 500 000 km 2 are Canada, Russia and Mexico. These are dominant soils in 

Curacao, Aruba, Bahamas, Bonaire, Saint Eustatius, Saba, Cayman Islands, Norway and El Salvador. 


587 


a 

b 

Photo by ISRIC World Soil Information Photo by ISRIC World Soil Information 

Figure A 31 (a) A Regosol profile and (b) the associated landscape, China. 


588 


ARENOSOLS (Psamments) 

Arenosols are sandy soils that may have diagnostic horizons below meter. These soils have low water 

holding capacity and where there is no plant cover they are subject to wind transport. Profile horizonation is 

A-C (Figure A 32). 

Arenosols occur in any climate except permafrost. Vegetation varies widely with climate. They are found 

on level to steep slopes. Typical textures are sandy or loamy sand on dunes, beaches, lacustrine deposits or 

weathered sandstone or coarse granite. The age can be recent to Pliocene or older. 

Arenesols occupy approximately 1 300 million ha or about 10 percent of the land surface of the globe. 

Countries with more than 400 000 km 2 are Australia, Sudan, China, Angola and Botswana. These are the 

dominant soils in Botswana and Angola. 


589 


Figure A 32 (a) An Arenosol profile in 

South Korea and (b) an Arenosol profile in 

New Mexico. 

Photo by H. Eswaran, USDA. 

a 

b 

Photo by H. Eswaran, USDA. 


590 


FLUVISOLS (Fluvents, Fluv-Subroups) 

Fluvisols are soils developedin fluvial, lacustrine or marine deposits A noted characteristic is an irregular 

decrease in organic carbon. In fact, significant soil organic carbon is buried by depositional events. Typical 

horizonation is A-C 2 

-C 3 

-Ab (Figure A 33). 

Fluvisols are found in all climates except permafrost. Vegetation depends on climate and proximity to 

water. The relief is usually level. Most of the soils are of recent origin. 

Fluvisols occupy approximately 350 million ha of the land surface of the globe. Countries with more than 

20 million ha are Russia, China and Indonesia. 


591 


Figure A 33 (a) A Fluvisol profile in 

Wisconsin and (b) a Fluvisol profile in 

Germany. 

Photo from the SSSA Marbut Slide Set 

a 

b 

Photo by H. Eswaran, USDA 


592 


9 | Permanently flooded soils 

WASSENTS, WASSISTS (subaquatic in Histosols and Fluvisols) 

Diagnostic horizons are typically absent from subaqueous soils. The exception is horizons formed from 

the accumulation of organic materials derived from submerged aquatic vegetation. Buried horizons have 

been observed because of sea level rise. In estuarine subaqueous soils, sulphides typically accumulate in low 

energy environments (sulphidization). Another important process is pedoturbation (faunal). This is especially 

the case in estuarine subaqueous soils where benthic organisms such as clams and worms burrow and mix 

the upper soil materials. 

Subaqueous soils are found in shallow areas of lakes, ponds and estuarine systems such as bays and 

lagoons in any climate. The distribution of the different subaqueous soil types typically follows the 

submerged landscape which is broken into different units such as submerged beach, bay bottom, washover 

fan, or flood-tidal delta. The parent materials are marine or lake sediments that have been brought in 

by streams and rivers emptying into the system or through inlets bringing in tidal water and sediment, or 

sediments brought in during storm events where over-wash events move materials from the barrier island 

into the lagoon. These are young soils, similar to floodplains in the subaerial system, and having little profile 

development. Buried horizons are common (figure A 34). 

Subaqueous soils provide the structure and habitat for the range of benthic organisms that live in these 

systems. Submerged aquatic vegetation is rooted in these soils and obtains some nutrients from the soils. 

Recent studies have shown that subaqueous soils store and sequester equivalent amounts of soil organic 

carbon as their subaerial counterparts. These soils serve as sinks for heavy metals and under certain 

conditions are important for water quality, storing N and providing denitrification. Shellfish aquaculture for 

species such as hard clams and oysters is a common practice on shallow estuarine subaqueous soils. 

A range of submerged aquatic vegetation can be found rooted in these soils, depending on the location, 

climate and water quality. Common species in estuarine systems include eelgrass (Zostera marina), turtle 

grass (Thalassia sp.), and widgeon grass (Ruppia sp.). In freshwater systems pondweed (Potamogeton sp.), 

watermilfoil (Myriophyllum sp.), and fanwort (Cabomba sp.) are commonly found. 


593 


Figure A 34 (a) A Wassent profile and 

(b) the associated landscape, the 

Netherlands. 

a 


b 



594 


Kastanozems 

Histosols 

Gypsisols 

Greyzems 

Gleysols 

Ferralsols 

Fluvisols 

Cambisols 

Calcisols 

Chernozems 

Anthrosols 

Arenosols 

Andosols Alisols 

Acrisols 

No data 

Luvisols Glaciers 

Leptosols 

No data 

Salt Flats 

Urban, mining 

Water 

Sand Dunes 

Rock outcrop 

Vertisols 

Solonetz 

Solonchaks 

Regosols 

Podzols 

Plinthosols 

Planosols 

Phaeozems 

Podzoluvisols 

Nitisols 

Lixisols 

Island 

Figure A 35: Global Soil Map of the World based on HWSD and FAO Revised Legend 

(Nachtergaele and Petri, 2008) 


595 


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Blume, H.P. & Leinweber, P. 2004. Plaggen Soils: landscape history, properties and classification. J. Plant 

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Boettinger, J. 2002. Calcification. In R. Lal, ed. Encyclopaedia of Soil Science , pp. 131-134. New York, Marcel 

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Brinkman, R. 1979. Ferrolysis, a soil-forming process in hydromorphic conditions. Agricututal Research Report 

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Charzyński, P., Bednarek, R., Hulisz, P. & Zawadzka, A. 2013. Soils within Toruń urban area. In P. Charzyński, 

P. Hulisz & R. Bednarek, eds. Technogenic soils of Poland, pp. 17–30. Toruń, Polish Society of Soil Science. 357 pp. 

Chesworth, W., ed. 2008. Encyclopedia of soil science. Springer Science & Business Media 

Chichagova, O.A. 1985. Radiocarbon Dating of Soil Humus. Moscow, Nauka Publ. 157 pp. [in Russian]. 

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Coulombe, C.E., Dixon, J.B. & Wilding, L.P. 1996b. Mineralogy and chemistry of Vertisols. Developments in 

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Glossary of technical terms 

Aerobic: a condition in which molecular oxygen is freely available (ISO, 2013). 

Anaerobic: descriptive of a condition in which molecular oxygen is not available (ISO, 2013). 

Available water capacity: soil water content useable by plants, based on the effective root penetration 

depth (ISO, 2013). 

Bare Soil: a land cover class that includes any geographic area dominated by natural abiotic surfaces (bare 

soil, sand, rocks, etc.) where the natural vegetation is absent or almost absent (covers less than 2 percent) 

(Latham et al., 2014). 

Biodegradation: physical and chemical breakdown of a substance by living organisms, mainly bacteria 

and/or fungi (ISO, 2013). 

Contaminant: substance or agent present in the soil as a result of human activity (ISO, 2013). 

Cropland: a land cover class that includes all cultivated herbaceous crops, woody crops and multiple and 

layered crops (Latham et al., 2014). 

Decomposition: breakdown of complex organic substances into simpler molecules or ions by physical, 

chemical and/or biological processes (ISO, 2013). 

Desertification: land degradation in arid, semi-arid and dry sub-humid areas resulting from various factors, 

including climatic variations and human activities (UNCCD, 2011). 

Drylands: tropical and temperate areas with an aridity index (annual rainfall/annual potential evaporation) 

of less than 0.65 (UNEP, 2005). 

Grassland: a land cover class that includes any geographic area dominated by natural herbaceous plants 

(grasslands, prairies, steppes and savannahs) with a cover of 10 percent or more, irrespective of different 

human and/or animal activities e.g. grazing, selective fire management (Latham et al., 2014). 

Habitat ecosystem functions: the ability of soil or soil materials to serve as a habitat for micro-organisms, 

plants, soil-living animals and their interactions (ISO, 2013). 

Humification: decomposition of organic material followed by a synthesis of humic substances (ISO, 2013). 

Land: terrestrial bio-productive system that comprises soil, vegetation, other biota, and the ecological and 

hydrological processes that operate within the system (UNCCD, 2011). 

Leaching: the dissolution and movement of dissolved substances by water (ISO, 2013). 

Mineralization: final stage of the biodegradation of organic matter or organic substances into carbon 

dioxide, water and hydrides, oxides or other mineral salts (ISO, 2013). 

Mitigation (of land degradation): an intervention intended to reduce ongoing degradation at a stage 

when degradation has already begun. The main aim here is to halt further degradation and to start improving 

resources and their functions (FAO, 2015). 

Parent material: The unconsolidated and more or less chemically weathered mineral or organic matter 

from which the solum of soils is developed by pedogenic processes (Soil Science Society of America, 2008). 

Particle size distribution: distribution of the soil mineral particles according to predefined classes of size 

(ISO, 2013). 

Status of the World’s Soil Resources | Main Report Glossary of technical terms 

599

Pedon: the smallest, three-dimensional unit at the surface of the earth that is considered as a soil. It forms 

a conceptual foundation for the study of soils as geographic entities (Hole and Campbell, 1985). 

pH-value: the negative logarithm (base 10) of the concentration of hydrogen ions, expressed in moles/l in 

aqueous solution and varying between 0 (extremely acid) to 14 (extremely alkaline) (ISO, 2013). 

Rehabilitation: action to restore soil already degraded to such an extent that the original use is no longer 

possible and the land has become practically unproductive. Generally, long term and often costly investments 

are needed to show any impact (FAO, 2015). 

Shrub-covered area: a land cover class that includes any geographical area dominated by natural shrubs 

having a cover of 10 percent or more (Latham et al., 2014). 

Soil: the upper layer of the Earth’s crust transformed by weathering and physical/chemical and biological 

processes. It is composed of mineral particles, organic matter, water, air and living organisms organized in 

genetic soil horizons (ISO, 2013). 

Soil degradation: the diminishing capacity of the soil to provide ecosystem goods and services as desired 

by its stakeholders (refined from FAO, 2015). 

Soil ecosystem functions: description of the significance of soils to humans and the environment. Examples 

are: (1) control of substance and energy cycles within ecosystems; (2) basis for the life of plants, animals and 

man; (3) basis for the stability of buildings and roads; (4) basis for agriculture and forestry; (5) carrier of genetic 

reservoir; (6) document of natural history; and (7) archaeological and paleo-ecological document (ISO, 2013). 

Soil health: the continued capacity of the soil to function as a vital living system, within ecosystem and 

land-use boundaries, to sustain biological productivity, promote the quality of air and water environments, 

and maintain plant, animal, and human health (Doran, Stamatiadis and Haberern, 2002). 

Soil organic carbon (SOC): a summarizing parameter including all of the carbon forms for dissolved (DOC: 

Dissolved Organic Carbon) and total organic compounds (TOC: Total Organic Carbon) in soils (ISO, 2013). 

Soil organic matter (SOM): matter consisting of plant and/or animal organic materials, and the conversion 

products of those materials in soils (ISO, 2013). 

Soil Processes: physical or reactive geochemical and biological processes which may attenuate, 

concentrate, immobilize, liberate, degrade or otherwise transform substances in soil (ISO, 2013). 

Soil quality: all current positive or negative properties with regard to soil utilization and soil functions (ISO, 

2013). 

Soil structure: the arrangement of soil particles in a variety of recognized shapes and sizes (ISO, 2013). 

Soil threats: see Box ‘Soil Threat Definitions’ 

Solum: comprises the surface layer and subsoil layers that have been altered by soil formation (Soil Survey 

Staff, 1993). 

Sparse vegetation: a land cover class that includes any geographic areas where the cover of natural 

vegetation is between 2 percent and 10 percent (Latham et al., 2014). 

Sustainable land management (SLM): the use of land resources, including soils, water, animals and plants 

for the production of goods to meet changing human needs while ensuring the long term productive potential 

of these resources and the maintenance of their environmental functions (UNCED, 1992). 

Sustainable soil management (SSM): sets of activities that maintain or enhance the supporting, 

provisioning, regulating and cultural services provided by soils without significantly impairing either the soil 

functions that enable those services or biodiversity (adapted from GSP, 2015). 


600

Topsoil: the upper part of a natural soil that is generally dark coloured and has a higher content of organic 

matter and nutrients when compared to the (mineral) horizons below. It excludes the litter layer (ISO, 2013). 

Tree-covered area: a land cover class that includes any geographic area dominated by natural tree plants 

with a cover of 10 percent or more. Areas planted with trees for afforestation purposes and forest plantations 

are included in this class (Latham et al., 2014). 

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601

Authors and affiliations 

Adams, Mary Beth 

Adhya Tapan, Kumar 

Agus, Fahmuddin 

Al Shankithi, Abdullah 

Alavi Panah, Sayed Kazem 

Alegre, Julio 

Aleman, Garcia 

Alfaro, Marta 

Alyabina, Irina 

Anderson, Chris 

Anjos, Lucia 

Arao, Tomohito 

Arrouays, Dominique 

Asakawa, Susumu 

Aulakh, Milkha Singh 

Ayuke, Frederick 

Badraoui, Mohamed 

Bai, Zhaohai 

Baldock, Jeff 

Balks, Megan 

Balyuk, Svyatoslav 

Bardgett, Richard 

Basiliko, Nathan 

Bationo, André 

Batkhishig, Ochirbat 

Bedard-Haughn, Angela 

Bielders, Charles 

Black, Helaina 

Bock, Michael 

Bockheim, James 

Bondeau, Alberte 

Brinkman, Robert 

Bristow, Keith 

USDA Forest Service, West Viriginia, USA. 

KIIT School of Biotechnology, Odisha, India. 

Indonesian Soil Research Institute, Indonesia. 

ICBA, United Arab Emirates. 

University of Teheran, Iran. 

National Agrarian University La Molina, Peru. 

Ministry of Agriculture, Cuba. 

INIA, Remehue, Chile. 

Lomonosov Moscow State University, Russian Federation 

Massey University, New Zealand. 

Federal Rural University of Rio de Janeiro, Brazil. 

National Institute for Agro-Environmental Science (NIAES), Japan 

National Institute for Agricultural Research, France. 

Nagoya University, Japan. 

Punjab Agricultural University, India. 

University of Nairobi, Kenya. 

National Agronomic Research Institute, Morocco. 

China Agriculture University, China. 

CSIRO, Australia. 

University of Waikato, New Zealand. 

National Institute of Soil Science, Ukraine 

Manchester University, UK. 

Laurentian University, Canada 

AGRA Kenya. (Burkina Faso). 

Academy of Sciences, Mongolia 

University of Saskatchewan, Canada. 

Catholic University Louvain, Belgium (USA) 

The James Hutton Institute, UK 

Agriculture and AgriFood Canada, Canada 

Wisconsin University, USA. 

Mediterranean Institute of Biodiversity and Ecology (IMBE), 

France 

Soil Science Society The Netherlands. 

CSIRO, Australia. 

Status of the World’s Soil Resources | Main Report Authors and affiliations 

602

Broll, Gabrielle 

Bruulsma, Tom 

Bunning, Sally 

Bustamante, Mercedes 

Camps Arbestain, Marta 

Caon, Lucrezia 

Carating, Rodel 

Cerkowniak, Darrel 

Charzynski, Przemyslaw 

Chude, Victor 

Clark, Joanna 

Clothier, Brent 

Coelho, Maurício Rizzato 

Colditz, Roland René 

Collins, Alison 

Comerma, Juan 

Compton, Jana 

Condron, Leo 

Corso, Maria Laura 

Cotrufo, Francesca 

Critchley, William 

Cruse, Richard 

da Silva, Manuela 

Dabney, Seth 

Daniels, Lee 

de Souza Dias, Moacir 

Dick, Warren 

Dos Santos Baptista, Isaurinda 

Drury, Craig 

El Mustafa El Sheikh, Ahmed Elsiddig 

Elder-Ratutokarua, Maria 

Elliott, Jane 

Espinosa Victoria, David 

Fendorf, Scott 

Ferreira, Gustavo 

Flanagan, Dennis 

Fraser, Tandra 

Gafurova, Laziza 

Osnabrück University, Germany. 

Paul Fixen, IPNI. Guelph Ontario, Canada. 

FAO, United Nations. (UK) 

University of Brasilia, Brazil. 


FAO, United Nations (Italy). 

Bureau of Soils and Water Management, Philippines 


Nicolaus Copernicus University, Poland. 

Ministry of Agriculture, Nigeria. 

Reading University, England, UK. 

Food Ind. Science Centre, New Zealand. 

EMBRAPA, Brazil 

Comision Nacional para el Conocimiento y Uso de la 

Biodiversidad (CONABIO) 

NLRC, New Zealand. 

Soil Science Society Venezuela, Venezuela. 

EPA, USA 

Lincoln University, New Zealand. 

Min. Environment, Argentina. 

Colorado State University, USA (Italy) 

VU Amsterdam, The Netherlands (UK) 

Iowa state University, USA. 

Joint Research Centre, Italy (Sweden/Brazil) 

USDA, USA. 

Virginia Tech , USA. 

Universidade Federal de Lavras, Brazil. 

Ohio State University, USA 

INIDA, Cape Verde. 


University of Khartoum, Sudan. 

Secretariat of the Pacific Community 

Saskatchwan University, Canada. 

Colegio de Postgraduados in Mexico 

Stanford University, USA. 

INIA, Urugay 

Purdue University, USA. 

GSBI, Colorado State University, USA (Can) 

National University, Tashkent, Uzbekistan. 


603

Gaistardo, Carlos Cruz 

Gardi, Ciro 

Gerasimova, Maria 

Govers, Gerard 

Grayson, Sue 

Griffiths, Robert 

Grundy, Mike 

Hakki Emrah, Erdogan 

Hamrouni, Heidi 

Hanly, James 

Harper, Richard 

Harrison, Rob 

Havlicek, Elena 

Hempel, Jon 

Henriquez, Carlos Roberto 

Hewitt, Allan 

Hiederer, Roland 

Hong, Suk Young 

House, Jo 

Huising, Jeroen 

Ibánez, Juan José 

Indraratne, Srimathie 

Jain, Atul 

Jefwa, Joyce 

Jung, Kangho 

Kadono, Atsunobu 

Kawahigashi, Masayuki 

Kelliher, Frank 

Kihara, Job 

Konyushkova, Maria 

Krasilnikov, Pavel 

Kuikman, Peter 

Kuziev, Ramazan 

Lai, Shawntine 

Lal, Rattan 

Lamers, John 

Lee, Dar-Yuan 

Lee, Seung Heon 

Lehmann, Johannes 

INEGI, Mexico 

University of Parma,Italy 

Moscow State University, Russian Federation. 

Catholic University Leuven, Belgium. 

University of British Columbia, Canada. 

Centre for Ecology & Hydrology, UK 

CSIRO, Australia 

Ministry Food, Agriculture and Livestock, Turkey 

Ministry of Agriculture, Tunisia 


Murdoch University, Perth, Australia 

University of Washington 

Université de Neufchatel.(CH) 

NRCS, Lincoln, USA. 

University of Costa Rica, Costa Rica. 

Landcare Research, New Zealand. 

Joint Research Center - EU (Germany) 

National Academy of Agricultural Science, RDA, South Korea 

University of Bristol, England, UK. 

TSBF-CIAT, Kenya (The Netherlands) 

Spanish National Research Council 

University of Peradeniya, Sri Lanka. 

Department of Atmospheric Sciences, University of Illinois, USA 

TSBF-CIAT, Kenya. 

National Academy of Agricultural Science, RDA, South Korea 

Tottori University of Environmental Studies, Japan 

Tokyo Metropolitan University , Japan 

AgResearch, New Zealand. 

CIAT, Kenya. 

Moscow State University, Moscow, Russia 

Moscow State University, Russian Federation. 

Wageningen University, The Netherlands. 

National Institute Soil Science, Uzbekistan. 

MWH Americas Inc. Taiwan branch. 

Ohio State University, USA 

Bonn University, Germany. 

National Taiwan University, Taiwan 

Korea Rural Community Corp., Korea 

Cornell University, USA (Germany) 


604

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Lobb, David 

Ma, Lin 

Macias, Felipe 

Maina, Fredah 

Mamo, Tekalign 

Mantel, Stephan 

McDowell, Richard 

McKenzie, Neil 

Medvedev, Vitaliy 

Mendonça-Santos, de Lourdes Maria 

Miyazaki, Tsuyushi 

Montanarella, Luca 

Moore, John 

Morrison, John 

Mubarak, Abdelrahman Abdalla 

Mung'atu, Joseph 

Muniz, Olegario 

Nachtergaele, Freddy 

Nanzyo, Masami 

Ndiaye, Déthié 

Neall, Vince 

Norbu, Chencho 

Noroozi, Ali Akbar 

Obst, Carl 

Ogle, Stephen 

Ogunkunle, Ayoade 

Okoth, Peter 

Omutu, Christian 

Or, Dani 

Owens, Phil 

Pan, Genxing 

Panagos, Panos 

Parikh, Sanjai 

Pasos Mabel, Susana (†) 

Paterson, Garry 

Paustian, Keith 

Pennock, Dan 

Pietragalla, Vanina 

NSW Office of Environment and Heritage, Australia 

University of Manitoba, Canada 

Institute of Genetic and Developmental Biology, CAS, China 

Universidade de Santiago de Compostela, Spain. 

Agricultural Research Institute, Kenya. 

Ministry of Agriculture, Ethiopia. 

World Soil Information (ISRIC), the Netherlands 

AgResearch, New Zealand. 


Ukrainan Agricultural Academy 

EMBRAPA, Brazil. 

University of Tokyo, Japan. 

Joint Research Center - EU (Italy) 

Colorado State University, USA 

University of Wollongong, Australia 

University of Khartoum, Sudan. 

Jomo Kenyatta University, Kenya. 

Soil Institute, Cuba. 

FAO, United Nations (Belgium) 

Tohoku University, Japan. 

CSE, Senegal 


Ministry of Agriculture and Forests, Buthan 

SCWM Institute, Iran. 

University of Melbourne, Australia 

Colorado State University, USA. 

University of Ibadan, Nigeria. 

TSBF-CIAT, Kenya. 

University of Nairobi, Kenya. 

ETH, Switzerland. 

Purdue University, USA. 

Nanjing Agricultural University, China. 

Joint Research Center - EU (Greece) 

University of California-Davis, USA. 

University Buenos Aires, Argentina. 

ISCW, Republic of South Africa. 

Colorado State University, USA. 

University of Saskatchewan, Canada. 

Min. Environment, Argentina. 


605

Pla Sentis, Ildefonso 

Polizzotto, Matthew 

Pugh, Thomas 

Qureshi, Asad 

Reddy, Obi 

Reid, D. Keith 

Reinsch, Thomas 

Richter, Dan 

Rivera-Ferre, Marta 

Robinson, David 

Rodriguez Lado, Luis 

Roskruge, Rick 

Rumpel, Cornelia 

Rys, Gerald 

Schipper, Louis 

Schoknecht, Noel 

Seneviratne, Sonia 

Shahid, Shabbir 

Sheffield, Justin 

Sheppard, Steve 

Sidhu, Gurjant 

Sigbert, Huber 

Smith, Pete 

Smith, Scott 

Sobocká, Jaroslava 

Sönmez, Bülent 

Spicer, Anne 

Sposito, Garrison 

Stolt, Mark 

Suarez, Don 

Taboada, Miguel 

Takata, Yusuke 

Tarnocai, Charles 

Tassinari, Diego 

Tien, Tran Minh 

Toth, Tibor 

Trumbore, Susan 

Tuller, Markus 

Universitat de Lleida, Spain (Venezuela) 

North Carolina State, USA. 

Karlsruhe Institute of Technology, Institute of Meteorology and 

Climate Research/Atmospheric Environmental Research (IMK- 

IFU), Germany 

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National Bureau of Soil Survey & Land Use Planning, India. 

Ontario Ministry of Agriculture, Canada 

World Soil Resources, USDA, USA. 

Duke University, USA. 

Universita de Vic, Spain. 

Centre for Ecology & Hydrology, UK. 

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INRA, France. 

Ministry Primary Industries, New Zealand 

The University of Waikato, New Zealand 

Department of Agriculture and Food, Western Australia 

ETH, Switzerland. 

International Centre for Biosaline Agriculture, UAE. 

Princeton University, USA (UK) 

ECOMatters Inc., Canada 

National Bureau of Soil Survey & Land Use Planning, India. 

Umweltbundesamt GmbH, Austria 

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Soil Science and Conservation Institute, Slovakia. 

Ministry of food agriculture and livestock, Turkey. 

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USDA Salinity Laboratory, Riverside, USA 

INTA, Argentina. 


Agriculture and Agri.Food, Canada. 

Universidade Federal de Lavras, Brazil. 

Soils and Fertilizers Research Institute, Vietnam. 

RISSAC, Hungary. 

Max Planck Institute, Germany (USA) 

University of Arizona, USA. 


606

Urquiaga Caballero, Segundo 

Urquiza Rodrigues, Nery 

Van Liedekerke, Marc 

Van Oost, Kristof 

Vargas, Rodrigo 

Vargas, Ronald 

Vela, Sebastian 

Vijarnsorn, Pisoot 

Vitaliy, Medvedev 

Vrscaj, Boris 

Wall, Diana 

Waswa, Boaz 

Watanabe, Kazuhiko 

Watmough, Shaun 

Webb, Mike 

Weerahewa, Jeevika 

West, Paul 

Wiese, Liesl 

Wilding, Larry 

Xu, Renkou 

Yagi, Kazuyuki 

Yan, Xiaoyuan 

Yemefack, Martin 

Yokoyama, Kazunari 

Zhang, Fusuo 

Zhang, Gan Lin 

Zhou, Dongme i 

Zobeck, Ted 

Embrapa Agrobiologica, Brazil. 

Ministry of Agriculture, Cuba. 

Joint Research Center, EU. (BEL) 

Catholic University Louvain, Belgium. 

CISECE, Mexico. 

FAO, United Nations (Bolivia) 

La Molina University, Peru. 

Chaipattana Foundation, Thailand. 

National Institute of Soil Science Ukraine. 

Agricultural Institute of Slovenia. 

Colorado State University, USA 

CIAT, Kenya. 

Hyogo Agricultural Institute, Japan. 

Trent University, Canada. 


University of Peradeniya, Sri Lanka. 

University of Minesota, USA. 

ARC - ISCW, Republic of South Africa. 

Texas A&M University, USA. 

Institute of Soil Science, CAS, China 


Institute of Soil Science, CAS, China 

IRAT/IITA, Cameroon 

National Agriculture and Food Research Organization (NARO), 

Japan 

China Agriculture University, China. 

Institute of Soil Science, China. 

Institute of Soil Science, CAS, China. 

USDA, USA. 


607

World’s Soil Resources

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