Advances in Agronomy, Volume 94

Agronomy D VA N C E S I N VOLUME 94 Advisory Board Paul M. Bertsch University of Georgia Ronald L. Phillips Univer...

Author: Donald L. Sparks

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Agronomy

D VA N C E S I N

VOLUME 94

Advisory Board Paul M. Bertsch University of Georgia

Ronald L. Phillips University of Minnesota

Kate M. Scow University of California, Davis

Larry P. Wilding Texas A&M University

Emeritus Advisory Board Members John S. Boyer University of Delaware

Kenneth J. Frey Iowa State University

Eugene J. Kamprath North Carolina State University

Martin Alexander Cornell University

Prepared in cooperation with the American Society of Agronomy, Crop Science Society of America, and Soil Science Society of America Book and Multimedia Publishing Committee David D. Baltensperger, Chair Lisa K. Al-Amoodi Kenneth A. Barbarick

Hari B. Krishnan Sally D. Logsdon Michel D. Ransom

Craig A. Roberts April L. Ulery

Agronomy D VA N C E S I N

VOLUME 94 Edited by

Donald L. Sparks Department of Plant and Soil Sciences University of Delaware Newark, Delaware

AMSTERDAM • BOSTON • HEIDELBERG • LONDON NEW YORK • OXFORD • PARIS • SAN DIEGO SAN FRANCISCO • SINGAPORE • SYDNEY • TOKYO Academic Press is an imprint of Elsevier

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Contents CONTRIBUTORS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . PREFACE . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

xi xiii

SOIL REDOX POTENTIAL: IMPORTANCE, FIELD MEASUREMENTS, AND OBSERVATIONS Sabine Fiedler, Michael J. Vepraskas and J. L. Richardson I. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . A. Redox and Wetland Issues . . . . . . . . . . . . . . . . . . . . . . . . . . . . . B. Relevance of Redox Measurements in Soil Science . . . . . . . . . . . II. Potentiometric Measuring Techniques . . . . . . . . . . . . . . . . . . . . . . . A. Reference Cell . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . B. Working Redox Electrodes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . C. Problems of Soil Redox Measurements. . . . . . . . . . . . . . . . . . . . D. Testing Electrodes Prior to Installation . . . . . . . . . . . . . . . . . . . III. Data Interpretation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . A. Correction of Field Data . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . B. Variability . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . C. Pooling of Long-Term Data Sets . . . . . . . . . . . . . . . . . . . . . . . . IV. Alternative Methods for Assessing Reduction in the Field . . . . . . . A. Iron-Coated (IRIS) Tubes. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . B. Dyes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . C. Zero Valence Iron Rods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . V. Field Installation and Procedures for Redox Potential Measurements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . A. Number of Electrodes to Install . . . . . . . . . . . . . . . . . . . . . . . . . B. Installing Pt Electrodes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . C. Reading the Electrodes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . D. Common Field Problems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . E. Field pH Measurements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . F. Interpreting Redox Potential. . . . . . . . . . . . . . . . . . . . . . . . . . . . VI. Summary . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

v

2 2 5 11 13 15 20 24 25 25 27 30 32 32 34 35 36 36 38 41 41 42 43 44 44

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INCREASING RICE PRODUCTION IN SUB-SAHARAN AFRICA: CHALLENGES AND OPPORTUNITIES V. Balasubramanian, M. Sie, R. J. Hijmans and K. Otsuka I. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . II. Rice Demand and Supply . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . III. Wetlands: The Potential Resource for Rice Production in SSA . . . . A. Definition, Area, and Distribution of Wetlands . . . . . . . . . . . . . B. Types and Characteristics of Wetlands . . . . . . . . . . . . . . . . . . . . IV. Rice Soil Resources . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . A. Dryland Soils and Their Characteristics . . . . . . . . . . . . . . . . . . . B. Wetland Soils and Their Characteristics . . . . . . . . . . . . . . . . . . . V. Agroclimatic Zones and Rice Ecosystems . . . . . . . . . . . . . . . . . . . . A. Dryland Rice Ecosystems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . B. Wetland Rice Ecosystems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . VI. Rice Production Constraints in SSA . . . . . . . . . . . . . . . . . . . . . . . . A. Physical, Biological, and Management Constraints . . . . . . . . . . B. Human Resource Constraints . . . . . . . . . . . . . . . . . . . . . . . . . . . C. Socioeconomic and Policy Constraints . . . . . . . . . . . . . . . . . . . . VII. Rice Research and Technology Development During the Past 20 Years . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . A. Rice Germplasm, Breeding, and Variety Development . . . . . . . . B. Rice Seed Production and Distribution Services . . . . . . . . . . . . . C. Crop Establishment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . D. Nutrient Management. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . E. Water Management for Rainfed and Irrigated Areas . . . . . . . . . F. Weeds, Insect Pests, and Diseases and Their Management . . . . . G. Grain Quality Management: From Breeding to Milling . . . . . . . H. Diversification of Rice Farming Systems . . . . . . . . . . . . . . . . . . I. ICM for Rice . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . VIII. Rice Intensification Issues and Thoughts for the Future . . . . . . . . . A. Rice Intensification in Relation to Vector-Borne Human Diseases . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . B. Environmental Issues Related to Rice Intensification in SSA . . . C. Preparing for the Impact of Climate Change . . . . . . . . . . . . . . . D. Technology Delivery and Deployment Issues . . . . . . . . . . . . . . . E. Policy Support for Rice Intensification in SSA . . . . . . . . . . . . . . IX. Conclusions: Challenges to and Opportunities for Enhancing Rice Production in SSA . . . . . . . . . . . . . . . . . . . . . . . . . Acknowledgments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

57 58 63 64 64 67 67 68 71 72 77 85 85 85 86 86 87 97 99 99 107 111 114 115 115 117 117 120 121 122 123 124 125 126

CONTENTS

vii

PHOSPHATE REACTION DYNAMICS IN SOILS AND SOIL COMPONENTS: A MULTISCALE APPROACH Yuji Arai and D. L. Sparks I. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . II. P Chemistry . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . III. Phosphate Adsorption on Soil Components. . . . . . . . . . . . . . . . . . . A. Phosphate Adsorption on Soils (Empirical Approaches) . . . . . . B. Phosphate Retention as AVected by Physicochemical Properties of Soils . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . C. pH EVects on Phosphate Adsorption on Variable Charge Minerals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . D. Phosphate Adsorption on Metal Oxides . . . . . . . . . . . . . . . . . . . E. Phosphate Adsorption on Phyllosilicate Minerals . . . . . . . . . . . . F. Temperature EVects on P Adsorption on Soil Components . . . . G. I EVects on P Surface Complexation . . . . . . . . . . . . . . . . . . . . . IV. Phosphate Surface Complexation on Soil Components . . . . . . . . . . A. Surface Complexation-Modeling Approaches . . . . . . . . . . . . . . . B. Electrophoretic Mobility Measurement Studies . . . . . . . . . . . . . C. Ex Situ Spectroscopic Studies . . . . . . . . . . . . . . . . . . . . . . . . . . . D. In Situ Spectroscopic Studies . . . . . . . . . . . . . . . . . . . . . . . . . . . V. Residence Time EVects on Phosphate Adsorption and Desorption in Soils and Soil Components . . . . . . . . . . . . . . . . . . . . A. Residence Time EVects Theory . . . . . . . . . . . . . . . . . . . . . . . . . . B. Slow Adsorption . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . C. Slow Desorption Process and Hysteresis . . . . . . . . . . . . . . . . . . . D. Solid-State, Inter-, and Intraparticle DiVusion . . . . . . . . . . . . . . E. Surface Precipitation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . F. Higher Energy Binding Through Chemical Reconfiguration. . . . VI. Future Research Needs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

136 137 138 138 139 140 141 142 143 145 146 146 147 148 149 159 159 160 161 165 167 170 171 171

ECOLOGICAL AGRICULTURE IN CHINA: PRINCIPLES AND APPLICATIONS Huixiao Wang, Longhua Qin, Linlin Huang and Lu Zhang I. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . A. Alternative Agriculture . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . B. Ecological Agriculture in the West . . . . . . . . . . . . . . . . . . . . . . . C. International Sustainable Agriculture . . . . . . . . . . . . . . . . . . . . .

182 183 183 184

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II. Chinese Ecological Agriculture . . . . . . . . . . . . . . . . . . . . . . . . . . . . A. Environmental Problems in China . . . . . . . . . . . . . . . . . . . . . . . B. Researches on SA in China . . . . . . . . . . . . . . . . . . . . . . . . . . . . C. Characteristics of CEA . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . D. Basic Principles of CEA . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . III. Development and Achievements of CEA . . . . . . . . . . . . . . . . . . . . . IV. Practical Aspects of CEA. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . A. Vertically Distributed Farming . . . . . . . . . . . . . . . . . . . . . . . . . . B. Multilevel Organic Substance Utilization . . . . . . . . . . . . . . . . . . C. Energy Exploitation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . D. Integrated Control Techniques . . . . . . . . . . . . . . . . . . . . . . . . . . E. Introducing New Varieties . . . . . . . . . . . . . . . . . . . . . . . . . . . . . F. Comprehensive Management of the Agricultural Environment . . V. Case Studies of CEA . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . A. Ecological Agriculture in Mountainous Regions . . . . . . . . . . . . B. Water-Collecting Ecological Agriculture in Western China . . . . C. From Ecological Agriculture to Ecological Industry . . . . . . . . . VI. Problems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . A. From Theory to Practice . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . B. Poverty . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . C. Lack of Systematic Theoretical Research . . . . . . . . . . . . . . . . . . D. Small Production Scale . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . E. Lack of Funds . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . F. Education of Farmers . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . G. Lack of Market Competiveness . . . . . . . . . . . . . . . . . . . . . . . . . VII. Concluding Remarks . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . A. Further Studies on Theories of CEA . . . . . . . . . . . . . . . . . . . . . B. Modern Techniques Application . . . . . . . . . . . . . . . . . . . . . . . . . C. Organic Food: Filling a Gap in the Market . . . . . . . . . . . . . . . . D. Ecological Agriculture and Township Enterprise . . . . . . . . . . . . E. International Cooperation in Ecological Agriculture . . . . . . . . . Acknowledgments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

185 185 186 187 188 191 193 193 193 194 194 194 195 195 195 197 199 202 202 202 203 203 203 203 204 204 204 204 205 205 205 206 206

COTTON LEAF PHOTOSYNTHESIS AND CARBON METABOLISM W. T. Pettigrew and T. J. Gerik I. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . II. Genetic Variability . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . A. CO2 Exchange Rate and Stomatal Conductance . . . . . . . . . . . . B. Chlorophyll Fluorescence . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

210 210 210 214

CONTENTS C. Photosynthetic Enzymes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . D. A:Ci Curves. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . E. 13C Discrimination . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

III. Management and Environmental EVects . . . . . . . . . . . . . . . . . . . . . A. Plant Growth Regulators . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . B. Plant Nutrition and Soil Fertility . . . . . . . . . . . . . . . . . . . . . . . . C. Moisture Stress . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . D. Temperature . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . IV. Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . V. Summary . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

ix 215 216 217 219 219 221 223 225 229 229 229

THE IMPACTS OF GRAZING ANIMALS ON THE QUALITY OF SOILS, VEGETATION, AND SURFACE WATERS IN INTENSIVELY MANAGED GRASSLANDS G. S. Bilotta, R. E. Brazier and P. M. Haygarth I. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . II. Impact of Treading by Grazing Animals on Grassland Soils . . . . . . III. Factors Influencing the Amount and Form of Soil Structural Alteration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . A. Animal Species and Age . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . B. Stocking Density . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . C. Soil Moisture . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . D. Vegetation Cover . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . IV. Forms of Soil Structural Alteration Resulting from Treading by Grazing Animals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . A. Soil Compaction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . B. Soil Pugging . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . C. Soil Poaching . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . V. The Impacts of Soil Structural Alteration by Grazing Animals . . . . A. Treading and Soil Physical Properties. . . . . . . . . . . . . . . . . . . . . B. Treading and Soil Hydrology . . . . . . . . . . . . . . . . . . . . . . . . . . . C. Treading and Vegetation Growth . . . . . . . . . . . . . . . . . . . . . . . . D. Treading and Soil Fauna . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . VI. The Impact of Defoliation by Grazing Animals on Grassland Vegetation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . A. Animal Species and Age . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . B. Stocking Density . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . C. Vegetation Response . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

238 240 241 241 242 243 244 245 245 246 247 249 249 251 252 254 254 255 256 257

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CONTENTS VII. Impact of Excretion by Grazing Animals on Vegetation, Soils, and Surface Waters in Intensively Managed Grasslands . . . . . . . . . A. Livestock Wastes as a Source of Nutrients . . . . . . . . . . . . . . . . . B. Livestock Wastes as a Source of Pathogens . . . . . . . . . . . . . . . . VIII. Impacts of Grazing Animals on the Water Quality of Surface Waters in Intensively Managed Grasslands . . . . . . . . . . . . . . . . . . . A. Soil Erosion and Sedimentation Problems . . . . . . . . . . . . . . . . . B. Eutrophication . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . C. Pathogenic Contamination . . . . . . . . . . . . . . . . . . . . . . . . . . . . . IX. Environmental Degradation by Grazing Animals: Recovery and Remediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . A. Natural Recovery of Soil Physical Condition Following Treading Damage . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . B. Mitigation and Damage Reduction Methods . . . . . . . . . . . . . . . X. Future Research . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Acknowledgments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

INDEX . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

258 258 259 260 262 263 265 266 266 266 272 273 273

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Contributors Numbers in parentheses indicate the pages on which the authors’ contributions begin.

Yuji Arai (135), Department of Entomology, Soils and Plant Sciences, Clemson University, Clemson, South Carolina 29634 V. Balasubramanian (55), International Rice Research Institute (IRRI), Metro Manila, Philippines G. S. Bilotta (237), Department of Geography, University of Exeter, Exeter, Devon EX4 4RJ, United Kingdom; Cross Institute Programme for Sustainable Soil Function (SoilCIP), Institute of Grassland and Environmental Research (IGER), North Wyke Research Station, Devon EX20 2SB, United Kingdom R. E. Brazier (237), Department of Geography, University of Exeter, Exeter, Devon EX4 4RJ, United Kingdom Sabine Fiedler (1), Institut fu¨r Bodenkunde und Standortslehre, Universita¨t Hohenheim, D-70593 Stuttgart, Germany T. J. Gerik (209), Blackland Research Center, Temple, Texas 76502 P. M. Haygarth (237), Cross Institute Programme for Sustainable Soil Function (SoilCIP), Institute of Grassland and Environmental Research (IGER), North Wyke Research Station, Devon EX20 2SB, United Kingdom R. J. Hijmans (55), International Rice Research Institute (IRRI), Metro Manila, Philippines Linlin Huang (181), Key Laboratory for Water and Sediment Sciences, Ministry of Education, College of Water Sciences, Beijing Normal University, Beijing 100875, People’s Republic of China K. Otsuka (55), Foundation for Advanced Studies on International Development (FASID) and National Graduate Institute for Policy Studies, 7-22-1 Roppongi, Minatoku, Tokyo 106-8677, Japan W. T. Pettigrew (209), United States Department of Agriculture, Agricultural Research Service, Crop Genetics and Production Research Unit, Stoneville, Mississippi 38776 Longhua Qin (181), Key Laboratory for Water and Sediment Sciences, Ministry of Education, College of Water Sciences, Beijing Normal University, Beijing 100875, People’s Republic of China J. L. Richardson (1), USDA-NRCS National Soil Survey Center, Lincoln, Nebraska 68508 M. Sie (55), Africa Rice Center (WARDA), 01 BP 2031 Cotonou, Benin, West Africa D. L. Sparks (135), Department of Plant and Soil Sciences, University of Delaware, Newark, Delaware 19717

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CONTRIBUTORS

Michael J. Vepraskas (1), Department of Soil Science, North Carolina State University, Raleigh, North Carolina 27695 Huixiao Wang (181), Key Laboratory for Water and Sediment Sciences, Ministry of Education, College of Water Sciences, Beijing Normal University, Beijing 100875, People’s Republic of China Lu Zhang (181), CSIRO Land and Water, Canberra, ACT 2601, Australia

Preface Volume 94 continues the rich tradition of Advances in Agronomy, publishing state-of-the-art reviews on timely and important topics in the plant and soil sciences. Chapter 1 is an excellent overview of soil redox potential, including discussions on redox and wetland issues, importance of redox measurements in soil science, techniques, limitations of redox measurements, field measurements, and data interpretation. Chapter 2 is a comprehensive review on rice production and productivity in Africa. Production trends, challenges, and opportunities as well as research frontiers are presented. Chapter 3 discusses advances in understanding phosphorus dynamics in soils, with an emphasis on the use of advanced analytical techniques such as in situ X-ray absorption and Fourier transform infrared spectroscopy to elucidate reaction mechanisms. Chapter 4 covers ecological agriculture in China and includes specific case studies. Chapter 5 reviews advances in understanding cotton leaf photosynthesis and carbon metabolism and impacts on cotton improvement and production. Chapter 6 reviews the impacts of grazing animals on soil quality, vegetation, and surface water quality. I am grateful to the authors for their outstanding contributions. DONALD L. SPARKS University of Delaware Newark, Delaware

xiii

SOIL REDOX POTENTIAL: IMPORTANCE, FIELD MEASUREMENTS, AND OBSERVATIONS Sabine Fiedler,1 Michael J. Vepraskas2 and J. L. Richardson3 1

Institut fu¨r Bodenkunde und Standortslehre, Universita¨t Hohenheim, D‐70593 Stuttgart, Germany 2 Department of Soil Science, North Carolina State University, Raleigh, North Carolina 27695 3 USDA‐NRCS National Soil Survey Center, Lincoln, Nebraska 68508

I. Introduction A. Redox and Wetland Issues B. Relevance of Redox Measurements in Soil Science II. Potentiometric Measuring Techniques A. Reference Cell B. Working Redox Electrodes C. Problems of Soil Redox Measurements D. Testing Electrodes Prior to Installation III. Data Interpretation A. Correction of Field Data B. Variability C. Pooling of Long‐Term Data Sets IV. Alternative Methods for Assessing Reduction in the Field A. Iron‐Coated (IRIS) Tubes B. Dyes C. Zero Valence Iron Rods V. Field Installation and Procedures for Redox Potential Measurements A. Number of Electrodes to Install B. Installing Pt Electrodes C. Reading the Electrodes D. Common Field Problems E. Field pH Measurements F. Interpreting Redox Potential VI. Summary References

Reduction and oxidation measurements create important data for analysis of wet soils. These measurements are actually recordings of voltage (EH) over time between a reference electrode and a sensor electrode inserted into 1 Advances in Agronomy, Volume 94 Copyright 2007, Elsevier Inc. All rights reserved. 0065-2113/07 $35.00 DOI: 10.1016/S0065-2113(06)94001-2

2

S. FIEDLER ET AL. a soil. The sensor electrodes are usually made of platinum wire (Pt electrode). Hydric soils require a period of reduction, and these measures can provide the length of time that the reduction process is occurring. The voltage results from an exchange of electrons between a redox couple such as ferrous and ferric iron during the process of reduction and oxidation. In soils that have fluctuating wet and dry conditions, wide fluctuations in Eh occur. Micro site differences complicate these measurements in that anaerobes may be active and at 1‐cm away they are completely inactive. The ferrous– ferric iron couple usually dominates these measurements but other couple often contributes complicating the measurements and interpretations of the data. Reference electrodes are often fine for laboratory work but are not rugged enough for the field. In this chapter, suggestions for improvement are discussed. Field‐measuring equipment and the Pt electrode are also sensitive and subject to problems that often lead to spurious results in the field. These problems are discussed at length. Iron‐coated tubes and other methods of establishing redox conditions are relatively recent and are # 2007, Elsevier Inc. discussed, also.

I. INTRODUCTION A. REDOX

AND

WETLAND ISSUES

Wetland environmental issues have increased in importance in the last three decades. The public became aware of the importance of wetlands after centuries of drainage and other degradation (Mitsch and Gosselink, 1986). In the United States, legal aspects of wetland preservation have increased the need for wetland delineation procedures. Hydric soils are defined as follows: ‘‘A hydric soil is formed under conditions of water saturation, flooding, or ponding long enough during the growing season to develop anaerobic conditions in the upper part’’ (Hurt et al., 2002). There is a need for improved assessment of anaerobic conditions, that is chemical reduction in soils, to identify hydric soils and wetlands. The classic measure of an anaerobic condition is the redox potential (Eh). This chapter is an overview and summary of the measurement of redox potential in soils, including equipment, possible problems, and alternatives. Hydric soils include soils that developed under suYciently wet conditions to support the growth and regeneration of hydrophytic vegetation. Soils that are saturated because of artificial measures are included in the concept of hydric soils. Also, soils in which the hydrology has been artificially modified are hydric if the soil was hydric in its unaltered state. Some soil series designated as hydric on hydric soil lists have phases that are not hydric because the water table is deep or because flooding and ponding do not

SOIL REDOX POTENTIAL

3

occur. Mottling, nodules, concretions, and other morphological features of soils associated with reduction and oxidation of Fe and Mn are summarized as redoximorphic features (Vepraskas, 1996) and called ‘‘paints of the earth’’ (Fanning and Fanning, 1989). A hydromorphic soil owes its morphology to water‐related processes, because the soils may be submersed below the water table continually, producing gley features or periodically creating redoximorphic features. These soils retain their morphology intact, or the morphology is only partly altered even when the soils are drained. When gleyed soils are drained, for instance, ferrous iron is oxidized to ferric iron on exposure to air (e.g., root channels) and therefore resembles redoximorphic features with the onset of drainage. In the United States, soils that are currently periodically or permanently water saturated are called ‘‘wetland soils’’ even if they are not actually in a ‘‘jurisdictional wetland’’ as defined by the US Army Corps of Engineers Manual (1987). Wetland soils do not diVer from hydromorphic soils, except that the hydromorphic soils may or may not be subjected to reduction and wetland soils are definitely reduced.

1.

Reduction and Oxidation (Redox) as a Process

Reduction, as a chemical process, occurs when an atom accepts an electron. This process increases the valence of an anion or decreases the valence of a cation. Oxidation is the reverse process and occurs when an atom loses an electron. In wetland soils, a fluctuating water table causes the Fe that coats soil particles to alternate between oxidized and reduced forms over the course of a year. Saturation for periods longer than a couple of weeks is usually suYcient to reduce Fe in the subsoil of many soils (Vepraskas, 2001). Vepraskas et al. (1999) created these conditions and observed that redoximorphic features and hydric soil field indicators could be created in less than 3 years with several wetting and drying cycles. Oxidation–reduction (redox) reactions in soils are mainly controlled by microbial activity and the presence of a supply of carbon for the microbes; during respiration, these organisms use organic substances as electron donors (Craft, 2001). Molecular oxygen acts as the preferred electron acceptor as long as there is a suYcient supply of oxygen. If the supply of oxygen is terminated, as is the case under conditions of saturation, the microbial activities switch from aerobic to facultative and eventually to anaerobic respiration. Once oxygen is consumed, alternative electron acceptors are used. The descending sequential order of acceptor preferences is NO 3 , MnO2, Fe(OH)3, , and CO (Bartlett, 1981, 1998; Bartlett and James, 1993). Soils containSO2 2 4 ing these elements provide an ecological niche for microbes; the microbes gain energy from the soil organic matter and are able to proliferate (Conrad and Frenzel, 2002). In either temporary or semipermanent water saturation

4

S. FIEDLER ET AL.

conditions, the diVusion of oxygen into soils is drastically curtailed, at least episodically. The biological and chemical processes are forced to change in an alternating manner with water saturation and then drainage. The soils change from an aerobic to an anaerobic environment and vice versa; soils that remain saturated (gley) are chronically reduced. These gley‐weathering zones are quite diVerent morphologically as seen in the field.

2.

Redox Potential

The abundance of oxidized and reduced chemical substances can be measured as a potential diVerence between an inert indicator electrode and a reference cell using a voltmeter or pH meter. The redox potential (Eh) is defined as the emf or potential of an electrode consisting of a redoxcouple (e.g., SO2 4 =H2 S) measured in a galvanic cell against the standard hydrogen electrode. Eh in soils generally ranges between 1 and þ1 V. This variation is due to the buVering eVect or poise of water on redox reactions (O2/H2O; H2O/H2) (Bartlett and James, 1993, 1995). Measured Eh must be regarded as an integrated parameter, which is triggered by the activity of living microbial communities. Thus, changes of external conditions, such as precipitation and water table (as indirect parameters of oxygen supply), temperature, and availability of organic matter, can all lead to changes in Eh values. Consequently, the redox potential can vary by several orders of magnitude both temporally (e.g., minutes, hours, or days) and spatially (e.g., micro sites, horizons, soil types, or landscapes) (Gao et al., 2002; Vepraskas et al., 1999). This enormous range of Eh values, even within a single soil horizon, makes the interpretation of redox data particularly complicated. According to Reddy et al. (2000), microbial aerobic activities reflect oxidizing conditions above an Eh of 300 mV; facultative reducing microbes are active from Eh 300 to 50 mV, or moderately reducing conditions. The preferred 4þ 3þ electron acceptors are first oxygen followed by NO 3 , Mn , and Fe . Obligate‐reducing microbes dominate at Eh levels below 50 mV. In these strongly reducing conditions, SO2 4 and CO2 are the usual electron acceptors. We use the Eh 300 mV as the break between aerobic and anaerobic conditions unless otherwise stated as per Reddy et al. (2000). In his classic paper on redox potential of marine sediments, ZoBell (1946) states, ‘‘The redox potential of sediments may be used advantageously in the study and interpretation of the morphology, general nature, and chemical processes in unconsolidated sediments.’’ ZoBell believes that the redox potential has a profound eVect on diagenesis of sediments, including conversion of organic matter into petroleum. Organic matter and mineral diagenesis are important redox‐related processes in soils. The term ‘‘transformation’’ is preferred as more general rather than diagenesis. As noted by ZoBell (1946),

SOIL REDOX POTENTIAL

5

Craft (2001), and Vepraskas (2001), an abundance of readily decomposable organic matter promotes reducing conditions. If organic matter is present in saturated soil, facultative reducing bacteria and perhaps allied organisms create the reducing conditions. The distribution (activity) of methanogenic organisms in anaerobic horizons is governed by fresh organic material. Carbon ‘‘hot spots’’ (Wachinger et al., 2000) are zones in which CH4 is produced because of very strong redox gradients that are caused by the porous structure of soil, possible colonies of microbes, concentrations of readily usable carbon, and the release of oxygen via the roots (Vepraskas et al., 1999). Bacteria in saturated soils that are consuming organic matter probably are the principal dynamic agents influencing redox potential. These bacteria produce fatty acids, such as acetic acid, ethanol, and other alcohols. Acetic acid and ethanol inhibit the growth of the bacteria and slow the consumption of organic matter over time. Organic matter undergoes a much slower transformation in anaerobic soils, especially once fermentation starts. As stated by Evans (2002): Under anaerobic conditions (lack of oxygen), pyruvic acid produced from glucose in the root cells of higher plants is converted to CO2 and ethyl alcohol. Ethyl alcohol tends to accumulate within root cells, and if anaerobic conditions persist, may accumulate at concentrations that can become toxic. Anaerobiosis creates an environment that impairs the growth rate of most plants that require oxygen for respiration. Drainage is the practice of removing excess water from land to minimize the occurrence of waterlogging. It has been an important management practice for centuries, and the consequences of poor drainage have been studied extensively over the past century. Therefore, wetland soils are often characterized by the accumulation of humus. As a result, wetland soils typically have much thicker organic horizons than their aerobic counterparts. The lower eYciency of the decomposition of organic matter under anaerobic conditions leads to an enrichment of water‐ soluble intermediate metabolites, which are characterized by high aromaticity and complexity of their molecules. In turn, a high level of solute organic reductants can act as an eYcient conveyor of redox‐sensitive elements (Lovley et al., 1998) and amplify the reducing conditions (Struyk and Sposito, 2001).

B. RELEVANCE

OF

REDOX MEASUREMENTS

IN

SOIL SCIENCE

The identification of redox intensities and dominant redox processes in soil under specific field conditions has a long tradition, starting with the pioneering work that was done more than 80 years ago (Bradfield et al., 1934;

S. FIEDLER ET AL.

6

Brown, 1934; Clark, 1925; Gillespie, 1920; Pearsall, 1938; Pearsall and Mortimer, 1939; Willis, 1932). Gillespie (1920) used platinum (Pt) electrodes to ascertain redox values, and Brown (1934) recognized that Eh was an intensity measure and introduced poise as a measure of redox capacity. Poise is the resistance to change in Eh when a small amount of oxidant removes electrons from a system or, conversely, a small amount of reductant adds electrons (Rowell, 1981). Poise and Eh are similar to buVer capacity and pH in soils. Poise increases with increasing concentration of oxidant and reductant (Bartlett and James, 1993; Brown, 1934; Rowell, 1981). Pearsall (1938) and Pearsall and Mortimer (1939) were able to relate various soil chemical properties and marsh vegetation zonation to redox conditions and measurements. Bradfield et al. (1934) used Eh measurements to locate suitable planting areas in orchards. After making initial Eh measurements on soil pastes or suspensions, it became apparent that it was only possible to detect oxidation–reduction conditions in soils when the measurements were carried out in situ (Bohn, 1968; Quispel, 1947; Willis, 1932). Quispel (1947), with some insight from very few field measures, related Eh (his rH) to water levels, moisture content, soil structure, and the amount of reducible materials. In situ measurements of Eh using temporary or permanently installed electrodes have routinely been applied in soil science since the 1960s (Aomine, 1962; Blume, 1968a; Mansfeldt, 2003; McKenzie et al., 1960; Meek et al., 1980; Thompson and Bell, 1996). Many important redox‐sensitive components, in particular trace elements, such as Se(IV)/Se(VI), and Cr(III)/Cr(VI), and associated elements (e.g., P, Mo, Si) undergo redox transformation (Runnells and Lindberg, 1990). Reactivity, mobility, toxicity, and bioavailability of these elements frequently depend on their redox state (Sigg, 2000). Toxic organic materials attenuate biomediation processes and are often activated or deactivated by redox processes (Doong and Wu, 1995). Furthermore, redox reactions trigger pedogenetic responses. These responses result in the field recognition of hydric soils from nonhydric soils in jurisdictional wetland delineations. In the following sections, brief illustrations are given of field redox measurements showing the importance of the standard procedures.

1.

Agriculture and Cultivation

Redox conditions are of basic importance to agriculture. Most of our agriculture systems are based on nonsaturated conditions. Anaerobiosis impairs the growth of most crops (notable exceptions are rice and cranberries). To minimize the negative condition of waterlogging, drainage has been the management tool of choice for many centuries. The consequences of

SOIL REDOX POTENTIAL

7

poor drainage have been heavily studied, but the reverse, or restoration of waterlogging, has not been studied as much. Wetland restoration and the use of wetlands for waste disposal and for improving water quality of agricultural runoV and for attenuating chemical contaminants, however, demonstrate the need for research.

2.

Plant Nutrients and Vitality

The decrease of the partial pressure of O2 leads to a decrease in the plants’ ability to take up mineral nutrients (Drew et al., 1988), depending on the minerals. For example, the uptake of potassium is hindered to a far larger degree than that of sodium (Drew et al., 1988). The reduced uptake of mineral nutrients and insuYcient O2 supply also impede root growth. Reduced root growth can result in lower production of growth regulators, such as hormones and other metabolites that are produced by roots, which in turn impacts the vitality of the entire plant. It is well recognized that oxygen deficiency inhibits synthesis of indolylacetic acid, gibberellins, and cytokinins by the root and increases the concentration of abscisic acid in the xylem sap (Reid and Bradford, 1984). The increase in free abscisic acid in leaves coincides with decreases in leaf water potential and turgidity (Santiago et al., 2000). Changes in root–shoot chemistry can also interfere with photosynthesis, ion transport (e.g., Kþ), and plant growth in general (Kludze and DeLaune, 1996). In most cases, the absence of oxygen leads to a greater mobility of redox‐ sensitive elements and hence to the accumulation of Mn2þ, Fe2þ, or sulfides up to phytotoxic levels (Pezeshki et al., 1988). Redox conditions are known to have a great influence on the transformation of nitrogen and basically determine its loss as N2O and N2 or its uptake as NHþ 4 or NO3 by plants. Furthermore, pedogenic Fe‐ and Mn‐(hydr‐)oxides are the most important adsorbents of micro‐ and macronutrients such as molybdenum (Gupta, 1997), copper (Cu) (Zhou and Wong, 2001), phosphorus (Frossard et al., 1995), and sulfate (Chao et al., 1962). Under anaerobic conditions, oxides dissolve and thus no longer contribute to nutrient adsorption. In contrast, sulfides are known to be eVective sinks for soluble metals in the soil solution by forming inaccessible compounds such as ZnS, CuS, FeS, or FeS2. 3. Heavy Metals During petroleum exploration and oil well preparation, drilling mud, crude oil, and saline ground water are spilled onto the surrounding soil near the well and often find their way into surface water (Al‐Sawari et al., 1998).

8

S. FIEDLER ET AL.

Apart from organic compounds (e.g., polycyclic aromatic hydrocarbons), these fluids contain heavy metals such as vanadium and nickel. Roth (2000) and Roth et al. (2000) have studied the vanadium migration at the oil production fields of the Agua Dulce District near Tabasco, Mexico. They observed that strong reducing conditions in organic‐enriched soils with Eh measurements ranging from 90 to 240 mV enhance the mobility of vanadium. They believe that vanadium can thereby enter into the human food chain. 4.

Toxic Organics

The degradation of toxic organics depends on their innate chemical properties and environmental conditions (Reddy et al., 2000). Recently, microbial reductive dechlorination of chlorinated ethane has become an important issue in the context of both natural attenuation and enhanced bioremediation. For example, the frequently observed groundwater contaminants tetrachloroethene (PCE) and trichloroethene (TCE) are dechlorinated under strictly anaerobic conditions by microorganisms (National Research Council, 1994). Dehalogenase enzymes are highly specific enzymes involved in this anaerobic dechlorination process; the highly chlorinated solvents are excellent electron acceptors. In addition, some of these bacteria, for example Dehalospirillum multivorans (Scholz‐Muramatsu et al., 1995), Dehalobacter restrictus (Holliger et al., 1998), and Dehalococcoides ethenogenes (Maymo´‐Gatell et al., 1997), are able to utilize these substances in their energy metabolism. Reductive dechlorination of PCE is based on the successive replacement of chlorine materials by hydrogen leading to TCE (Cirpka et al., 1999). Less chlorinated aliphatic hydrocarbons, such as cis‐1,2‐dichloroethene (DCE), or vinyl chloride, can be metabolized aerobically to CO2, chloride, and water by methane‐ and ethane‐oxidizing bacteria (Wackett, 1995). Depending on the environmental conditions and the presence of the dechlorinating bacteria, these processes can be utilized for groundwater remediation. In addition, anaerobic conditions are likely to cause an increase of soluble humic substances in wetland soils, which in turn enhances the solubility of toxic organics (Pardue et al., 1993). It is important to have fast, reliable field tests of redox to measure these conditions.

5.

Pesticides

There are numerous pesticides for which degradation is favored by anaerobic/reducing conditions (Seybold et al., 2001). We are using only a selected

SOIL REDOX POTENTIAL

9

few here. The degradation of such pesticides as 2,4,5‐trichlorophenoxyacetic acid (Gibson and Suflita, 1990), pentachlorophenol (Chang et al., 1996), and chloroanilines acid (Kuhn et al., 1990) in soils and aquifers is favored by strong reducing conditions (Eh < 100 mV). Seybold et al. (2001) observed rapid degradation of atrazine and metolachlor. Within 25 days in a strongly anaerobic condition, most of the pesticides had degraded to metabolites. The metabolites were not considered hazardous.

6.

Radionuclides

Groundwater originating from uranium mine mill‐tailing sites is often contaminated with uranium (Abdelouas et al., 1998, 2000; Buck et al., 1996). The behavior of such high‐level nuclear water is determined by oxic/anoxic conditions (Casas et al., 1998). The microbially mediated reduction of soluble U6þ to insoluble U4þ was proposed as a mechanism that removes dissolved uranium from waste streams. The potential of reductive precipitation was demonstrated with anaerobic microbes, such as iron‐ and sulfate‐reducing bacteria (Robinson et al., 1998).

7.

Methane

Methane, as part of the greenhouse gas problem, causes a global warming concern. Methane production from wetlands, therefore, needs to be addressed as an important issue. The emission of CH4 from wetland soils is the result of an interaction between methane production and oxidation. In anaerobic, gray soils (e.g., the Bg horizon in a Histic Humaquept), CH4 is produced by methanogenic bacteria (Fig. 1). CH4 can be oxidized by methanotrophic bacteria (consumption) (e.g., the Bg horizon during a dry period in an Aeric Endoaquept) on diVusion through the aerobic zone. Small areas of wet depressions, which are influenced by alluvial and colluvial processes (Endoaquepts), are eVective CH4 emitters (Sommer and Fiedler, 2002). Well‐aerated soils can act as eVective CH4 sinks (Smith et al., 2003). Therefore, the spatial arrangement of aerobic and anaerobic soils and of interhorizonal areas in individual soil profiles or even in pedons on landscapes is certainly a key factor for greenhouse gas atmospheric discharges (Fig. 1). The alternations of source and sink of these gases can be observed from the soil morphology itself (Fiedler and Sommer, 2000; Fiedler et al., 2005). An inexpensive assessment of this problem of methane as a greenhouse gas can be made by simply observing soil morphology. Figure 1 illustrates the hydrosequence described by Fiedler and Sommer (2000).

S. FIEDLER ET AL.

C Typic eutrudept Brownish Fe-enriched Bw-horizon

CO2 + CH4

Bg

Aeric endoaquept O or Ahorizon carbonenriched

>−75 mV

O

Bg

Production

BW

−Eh+

CH4

Anaerobic

CH3COOH

Aerobic

CO2 + 4H2

Anaerobic

H>>−75 mV

BW

Consumption

Transport

Aerobic

A CH4 + 2H2O

−Eh+ >−75 mV

CH4

<−75 mV

CH4 −Eh+

Emission Production Consumption

Emission

Uptake

<−75 mV

10

Histic humaquept Seasonal low water table

Blue-green Fe-depleted Gley-horizon

Figure 1 Schematic depiction of the relationship between CH4 flux, water table, redox conditions, soil morphology, and soil types.

The relationship of the water table to soil morphology, the redox potential, and the production and consumption of organic materials is important.

8.

Soil Genesis

Redox processes are regarded as primary mechanisms in the weathering of rocks and minerals (Walker, 1949). Redox conditions determine important properties of clay minerals such as cation‐exchange capacity (Stucki and Roth, 1977), surface area, and swelling behavior (Gates et al., 1993; Shen et al., 1992). Of all the elements comprising the crystal structures of minerals, Fe is one of the most interesting because it may be oxidized or reduced in situ (Stucki et al., 1987). In general, Fe occurs in its reduced form in primary minerals. Structural Fe2þ may be oxidized under aerobic conditions and cause an imbalance in the overall electrostatic charge of the crystal structure (Churchman, 2000). The redox status of Fe in crystal structures (Fe2þ ↔ Fe3þ) can (1) aVect the availability of nutrients (Komadel et al., 1995; Stucki et al., 1984) and (2) initiate clay mineral transformations in the course of pedogenesis following a sequence illite–vermiculite–smectite (Niederbudde and Fischer, 1980; Wilson, 1999). It is well known that this oxidation process leads to ferric Fe oxides (Schwertmann and Taylor, 1989). Stucki et al. (1987) found that octahedral Fe3þ in the crystal structures of smectites was reduced to Fe2þ by microorganisms indigenous to the soil.

SOIL REDOX POTENTIAL

11

Ernstsen et al. (1998) studied microbial reduction of structural Fe in minerals in comparison to chemical reduction. They found that the presence of microbes led to an increase of structural Fe2þ from 10 to 34% of the total Fe content in clay. However, the observed concentrations were lower than those obtained for chemically reduced (Na‐dithionite) subsamples (76–79%). The formation of redoximorphic features, such as Fe/Mn concentrations and nodules, is due to element redistribution (Blume, 1968b,c; Fiedler et al., 2004a; Richardson and Daniels, 1993; Richardson and Hole, 1979; Vepraskas et al., 1999) and has been reviewed by Vepraskas (1996, 2001). The development of depletion and accumulation horizons (Blume and Schlichting, 1985; Dobos et al., 1990; van Schuylenborgh, 1973) and bog iron ores (Blume, 1968b; Dobos et al., 1990; Kaczorek and Sommer, 2003) is also covered by Vepraskas (1996, 2001) and will be discussed later. The genesis of acid sulfate soils is mainly influenced by redox‐related processes (Fanning and Fanning, 1989). The exposure of the sulfide in these soils to air by soil disturbance, especially drainage, can lead to the generation of sulfuric acid and a change in carbonate concentration and pH and the formation of either gypsum or jarosite. Under anaerobic conditions, sulfates originating from seawater are reduced and nearly insoluble Fe‐ sulfides accumulate. On exposure to air, the sulfide oxidizes to sulfuric acid, which ‘‘destroys’’ the carbonates. Often the carbonates neutralize some of the acid and form gypsum. If few or no carbonates are present, the acid sulfate reactions form jarosite. A surplus of sulfuric acid leads to complete decalcification and finally to very strong acidification (pH < 3) (Fanning and Fanning, 1989).

II. POTENTIOMETRIC MEASURING TECHNIQUES The minimum setup for the measurement of Eh requires a working electrode and a reference electrode that are connected to a voltmeter or pH meter (Pearsall and Mortimer, 1939). The potential is transduced by an amplifier (included in the mV or pH meter) and transmitted to a data acquisition unit, such as a data logger (Flessa and Fischer, 1992), personal computer (PC) (Fiedler and Fischer, 1994), or write recorder (Le Brusq et al., 1987), via high‐quality, well‐protected, and insulated cables. Figure 2 shows a schematic arrangement of a general field measurement circuit using a galvanometer, an amplifier, two known resistances in the instrument, and two electrodes: a reference cell electrode and a working Pt electrode in the soil. This arrangement detects voltage in the soil (Eh) between the electrodes.

S. FIEDLER ET AL.

12 Connect reference electrode to "common" terminal

160 mv

C

Connect Pt electrode to voltage terminal

V

Voltmeter

Reference electrode, ceramic tip in contact with soil

Pt electrode is surrounded by a seal made from the soil extracted and pored into the hole as a slurry

Soil

Figure 2 Basic setup for the measurement of soil redox potentials, which is associated with a number of shortcomings (discussed in the text). For simplicity, only one redox electrode is shown, although multiple electrodes should be installed.

Reprocessing the signal impedance by means of an amplifier is essential because the potentials (DEhmax ¼ 2000 mV) are loaded with an extremely weak current (<10–7 A) (Morris and Stumm, 1967). To prevent voltage deterioration, the amplifier input impedance should be >108 O (preferably 1012 O). Although wetland science and soil science provide abundant opportunity for the application of Eh measurements, currently very few commercial electrodes that solve the voltage deterioration and low wattage problems are available for field use to evaluate soil redox conditions. The high costs of commercial electrodes have driven many researchers to design their own systems and have led to a large number of detailed publications on electrode systems (Blanchar and Marshall, 1981; Fischer and Schaller, 1980; Mann and Stolzy, 1972; Pang and Zhang, 1998; Quispel, 1947; Teasdale et al., 1998; Zhang and Pang, 1999). A summary of Eh electrode systems that have worked well in field practice for a variety of purposes is included. Technical problems of Eh electrode construction are discussed. A construction ‘‘manual’’ of currently used electrodes will be presented.

SOIL REDOX POTENTIAL

13

A. REFERENCE CELL Reference cells should be able to provide a defined and constant virtual grounding potential that must be stable against changes in outer electrolyte composition. Ives and Janz (1961) outline most of the diverse variety of reference electrodes and construction. A common reference cell is the Ag/AgCl single‐junction reference electrode. Most commercial versions consist of a glass tube containing 4‐M solution of KCl saturated with AgCl. The lower end of the electrode, which is placed in the soil, is sealed with a porous ceramic frit that acts as a diaphragm. The porous ceramic material allows the slow passage of the electrolyte in the electrode to the soil. The electrolyte forms a liquid junction with the soil or soil water. Within the electrode, an Ag wire is coated with a layer of AgCl. This wire dipped in the electrolyte solution is coupled or joined to a low‐noise cable, which connects to the pH/Eh meter. Generally, industrial electrodes have very small contact areas, or diaphragms, which may cause problems with the galvanic contact. Frequently, such reference electrodes in the field become erratic during the summer season due to a low moisture content. Also, the glass electrode body is fragile; a sturdier material such as plastic is suggested. Farrell et al. (1991) also noted that the small storage capacity of the KCl electrolyte in the commercial electrodes requires frequent refilling with KCl solution. Such a problem can hinder long‐term field use in remote areas and may add to project costs. Farrell et al. (1991) provide instructions for manufacturing reference cells with large diaphragms. Following these instructions with modifications, Fiedler et al. (2003) constructed a reference cell with a diaphragm that was 70 times larger than that of conventional reference cells (Fig. 3). These reference electrodes are able to stand up better in the field than commercial types. These reference cells are conventional Ag/AgCl electrodes designed with tough electrode parts wherever they come in contact with the soil. In contrast to this, however, the diaphragm consisted of a ceramic ring that had been cut from the head of a micro‐tensiometer. The body of the electrode consists of a polyvinyl chloride (PVC) shaft, which is closed by a hollow polyethylene plug at the top. At the bottom, the diaphragm is inserted via waterproof resin into the PVC shaft. A coiled Ag wire (99.9% Ag, 0.5 mm) was electrolyzed as follows: after the wire was cleaned in ethanol, the Ag wire (anode) and a Pt wire (cathode) were immersed in an HCl solution (1 M) and a potential of 1.7 V was set up for electrolysis. The process was continued for about 5 min. The formation of a gray‐brown coating should occur. After the electrolysis process, the Ag wire was cleaned under running water and left in water overnight. Then the

S. FIEDLER ET AL.

14

Figure 3

Reference electrode (see text for details).

Ag wire was soldered to the conductance Cu wire at the Ag/Cu junction and suspended into a perforated plexiglass tube (6‐mm i.d.), which was inserted into the PVC shaft through the plug. A recess in the interior of the plug contained the soldered joint and a part of the connecting cable to the Eh meter that was sealed with epoxy resin. The reference was filled with 3‐M KCl to which 1 g liter1 of AgCl (saturation) was added in order to prevent dissolution of the AgCl wire in the concentrated Cl solution. The electrode potential for this half‐cell is 204.6 mV relative to the standard hydrogen electrode at 25 C. The constructed reference cell showed a current of 2 10–5 A, whereas a conventional reference cell exhibited a current of 1 10–7 A. It was concluded that the galvanic contact could be maintained for a much longer time when a diaphragm with a larger diameter was used. However, excessive eZux of the electrolyte could be a source of ‘‘potassium fertilizer’’ and might be undesirable in some cases. Matia et al. (1991) tested Calomel (Hg2Cl2/Hg) and Ag/AgCl reference electrodes with single, double, ceramic, and sleeve junction types. They demonstrated that the type of junction influenced precision of the measurements. They observed that the highest precision occurred with sleeved junctions on Ag/AgCl electrodes. Some researchers recommended a KCl salt bridge for rapid and accurate measurements in soils with a low moisture content (Hostettler, 1984; Olness et al., 1989; Veneman and Pickering, 1983) or chemically severe milieu, such as sulfide‐enriched environments (Mansfeldt, 2003), or in acid sulfate soils (Charoenchamratcheep et al., 1987). According to Galster as quoted by

SOIL REDOX POTENTIAL

15

Frevert (1984), reactions between internal electrolyte and components of the soil solution can create deviations of up to 60 mV. Strong reducing conditions may liberate H2S, which then diVuses into the Ag/AgCl reference cell, impairing its function by coating the cell with AgS and hence plugging the diaphragm.

B. WORKING REDOX ELECTRODES A working field redox electrode must be chemically inert and must be an excellent electrical conductor. There are two general classes of sensor materials for redox electrodes: (1) metals and (2) semiconductors. The most important characteristics of such material can be summarized as follows: the electrode should allow an electron transfer between the oxidized and the reduced species of a redox couple through the electrode interface, without any participation or change of the electrode surface in the reaction (PfeiVer, 2000). Additionally any material used should have no catalytic eVect on the equilibrium (Galster, 2000). 1.

Metals as Electrode Sensors

No solid material is absolutely chemically inert. Galster (2000) states that a metal can be labeled as ‘‘inert’’ if that metal’s standard potential is more than 100 mV higher than the redox potential of the target sample. Frequently, metals from Group VIIIB, including iridium (Ir), platinum (Pt), palladium (Pd), and rhodium (Rh) (Pang and Zhang, 1998), and Group IB, including gold (Cater and Silver, 1961; Grenthe et al., 1992), are applicable in soils with redox potentials between 400 and 850 mV at pH 7. The most commonly used electrode material for Eh measurements is Pt (Biddle et al., 1995; Bohn, 1968, 1969; Galster, 2000; Matia et al., 1991). Pt oVers a high exchange current (10–3 mA cm2) and so can respond to a potential change. Pt also has a high standard potential and is therefore relatively inert: (Eo Pt/Pt2þ ¼ 1200 mV at 25 C). Enhanced durability is provided by employing a Pt 10% iridium electrode for added mechanical strength, which ensures a long period of trouble‐free service. In well‐drained soils, the Eh depends primarily on the presence of oxygen. The reduction of gaseous O2 is very slow and cannot be catalyzed by bright Pt. The potentials measured with electrodes of bright Pt are much too low and are poorly reproducible (Quispel, 1947). The electron transfer is enhanced by modifying the surface of the Pt sensor by platinization as recommended by Quispel (1947). We suggest the following as an example of the process. Briefly, clean the Pt wire with HNO3 and acetone. Dissolve 3 g of crystallized H2PtCl6 and 0.02 g of lead acetate in 100 ml of deionized water. In this

S. FIEDLER ET AL.

16

solution, submerge two Pt wires. Establish a potential of 3 V between the wires, and reverse the polarity of the potential every minute for 10 min. A slight production of gas should be observed. At this point, the process of electrolysis should be terminated. Submerge the wires briefly in a dilute solution of H2SO4, which has a generated voltage of about 4 or 5 V. The impurities of the Pt solution should be chemically removed with these steps. Finally, wash the Pt sensor in water.

2.

Semiconductors as Electrode Sensors

The most common nonmetal semiconductors are wax‐impregnated graphite (Biddle et al., 1995; Grundl and MacAlady, 1989) and glassy carbon (Grenthe et al., 1992; Matia et al., 1991; Teasdale et al., 1998). Generally, precision is lower in carbon‐fiber electrodes when contrasted to Pt electrodes. Stabilization times are similar, however, to those for Pt electrodes (Matia et al., 1991). Teasdale et al. (1998) found that the Pt electrodes were preferred over glassy carbon electrodes for applications in natural waters and soil. The glassy carbon electrodes were preferred for measuring redox potentials involving organic species. These electrodes indicate lower potentials than Pt electrodes in the same environment because carbon‐fiber had a diVering response to exactly the same redox systems as the Pt electrodes, requiring diVering interpretations (Matia et al., 1991). The lower cost graphite electrodes are easily constructed (Biddle et al., 1995) as follows. Graphite can be obtained from the anode of carbon batteries and cleaned using aqua regia (1:3 conc. HNO 3:conc. HCl). In order to solder an electrical contact lead to the graphite electrode, a Cu plating on a portion of the graphite is needed. This can be achieved by immersing a 1‐cm length of graphite in CuSO4 solution and connecting the electrode to a voltage supply with cathode polarity using a length of Cu as the anode. The uniformity of the Cu coating is improved by sanding the graphite with fine corundum paper. The optimum plating condition for a uniform coating of Cu is 1–5 V at 300 mA. With a 100‐fold less exchange current and lower precision, wax‐ impregnated graphite electrodes are less desirable than Pt electrodes for making Eh measurements, which are erratic even under the best conditions. Biddle et al. (1995) pointed out the impracticality of graphite under aerobic conditions. In aerobic conditions, graphite reacted with oxygen (Cgraphite þ O2 ↔ CO2 þ 4e). As a consequence, accumulation of electrons at the surface of the electrode will cause erroneously lower Eh readings than the actual Eh of the soil.

SOIL REDOX POTENTIAL

3.

17

Design and Construction

Schaller and Fischer (1981) examined the question of how much surface area should be exposed from an electrode. They found that surface area exposure was not a factor in their study. Commonly, an eVective length of 10 mm is used (Bailey and Beauchamp, 1971; DeLaune et al., 1983), although Yamane and Sato (1970) recommended 15 mm as a minimum. DeLaune et al. (1983) state, ‘‘Redox potential (Eh) measurements made in situ with Pt electrodes constructed by welding 1.5 cm of 1‐mm diameter Pt wire to 3‐mm diameter solid Cu wire, covering with insulation and sealing with epoxy resin so that only bare Pt was in contact with the sediment. The maximum diameter of the electrode was 5 mm.’’ This is illustrated in Fig. 4. Loss of current from soil solution to the Pt wire–Cu junction is a common problem with permanently installed electrodes. Usually there are water leaks through the epoxy resin, which result in erroneous and/or erratic measurements. Therefore, watertight or leakproof electrodes are essential if meaningful and reproducible results are to be obtained (Mann and Stolzy, 1972; Mueller et al., 1985). Watertight or leakproof electrodes can provide suYcient isolation of the junctions. Cogger et al. (1992) stated that fast‐curing epoxies might absorb water from wet soil, leading to deterioration of the epoxy and shorting the electrode. As a result, Eh measures in a soil can drop several months after the electrodes were installed due to connection problems and not field redox conditions. To circumvent the epoxy problems, Fiedler et al. (2003) recommended the application of binary silicon resin (SemicosilÒ , Wacker, Germany), which retains its viscosity after polymerization and does not suVer from capillary ruptures and water absorption (Fig. 4).

Figure 4 A basic Pt electrode structure with a copper connecting wire attached by heat shrink insulator material or a binary silicon resin. Past history of epoxy resins suggests extreme caution because of water absorption and short life in use.

S. FIEDLER ET AL.

18

Heat‐shrinkable electrical insulation material has also been used (Mann and Stolzy, 1972; Whisler et al., 1974; Young et al., 1979). Leak‐proof electrodes have also been constructed with dental ceramics (Pfisterer and Gribbohm, 1989), by melting of Pt and glass (Bohn, 1968; Whisler et al., 1974), and by shielding the metal wire with a heat‐shrinkable electrical insulation material (Mann and Stolzy, 1972; Young et al., 1979). Because of their fast reaction time, resins are the most frequently applied remedies (Mueller et al., 1985; Whisler et al., 1974). In some previous investigations, mercury was preferred as linkage between sensor and signal line (Austin and Huddleston, 1999; Bailey and Beauchamp, 1971; Ponnamperuma, 1972). Because of the potential for mercury contamination in the environment, this technique is not recommended. Commonly, the Pt junction is inserted at the end of a tube consisting of several materials such as PVC (Olness et al., 1989), tygon (Mann and Stolzy, 1972), stainless steel (Norrstro¨m, 1994), or glass (Mann and Stolzy, 1972).

4. Pretreatment Contaminants and surface coatings must be removed from electrodes for accurate Eh measurements. Numerous methods have been used for these pretreatments: Mechanical: Polished with a fine grade of alumina (Al2O3), CeO2, or diamante power (Galster, 2000; Schaller and Fischer, 1981). Chemical: In most cases, immersion in hydrochloric acid and subsequent washing with distilled water is a suitable cleaning technique (Galster, 2000; Ponnamperuma, 1972). Thermal: Grundl and MacAlady (1989) and Grundl (1994) recommended soaking the electrode in 6‐M HCl and heating in a flame to incandescence. In order to get reproducible values, Yamane and Sato (1970) recommended washing the electrode well with running, deionized water and burning it red with an ethanol or methanol flame. Bailey and Beauchamp (1971) cleaned their electrodes by first heating and cooling, followed by washing in a solution of 10% HCl plus 10% detergent.

5.

Testing Procedure

Eh electrodes need to be tested with known redox solutions for electrode reliability before installation. Such test solutions are used to identify gross malfunctions of the electrodes or the electronics used. Abnormalities can happen even with care in construction or from commercial equipment (Table I). Test solutions create a redox equilibrium, which produces a stable,

SOIL REDOX POTENTIAL

19

reproducible poised potential. Electrodes that vary substantially from the reference solution need cleaning or need to be reconstructed. Several of the common solutions used to test the accuracy of newly constructed electrodes are listed in Table I. ZoBell’s solution is commonly used for testing Pt electrodes. This solution is well known and generally accepted. Its temperature range is given in Table II. Test solutions are highly poised by design, which means that they may not perform adequately in low‐poise field conditions. Schaller and Fischer (1981) and Teasdale et al. (1998) suggest testing in both high‐ and low‐poise test solutions. Teasdale et al. (1998) had success with a simple 100:1 dilution of ZoBell’s solution. A suspension of pure quinhydrone in potassium acid phthalate also is commonly used. Quinhydrone solution becomes unstable above 30 C. Table I Various BuVer or Poise (Test) Solutions Suggested for Calibration of Working Pt Electrodes Potential versus standard hydrogen electrode (25 C)

Test solution ZoBells solution 3.3 10–3 M K4Fe(CN)6 þ 3.3 10–3 M K3Fe(CN)6 in 0.1‐M KCl Saturated quinhydrone in 0.05‐M potassium biphthalate (pH ¼ 4.01) Saturated quinhydrone in 0.05‐M potassium biphthalate (pH ¼ 7) Ferrous ammonium sulfate‐ferric ammonium sulfate solution 15%Ti(III)Cl3 in 0.2‐M sodium citrate a

References

þ430 mV

Teasdale et al. (1998); ZoBell (1946)

þ462 mV

Austin and Huddleston (1999); Mansfeldt (1993) Austin and Huddleston (1999) Light (1972)

480 mV

Zehnder and Wuhrmanna (1976)

Zehnder and Wuhrmann (1976) quoted by Teasdale et al. (1998).

Table II Eh of ZoBell’s Solution as a Function of Temperature (from Nordstrom and Wilde, 1998) Temperature ( C) 10 14 18 22 26

Eh (mV)

Temperature ( C)

Eh (mV)

467 457 448 438 428

30 34 38 40

418 407 397 393

S. FIEDLER ET AL.

20

Additionally, the Eh of a quinhydrone solution is pH dependent. Light (1972) suggests that electrodes can also be tested in an Fe(II)–Fe(III) solution, but this solution, like all other standard redox solutions, is so strongly poised that even unsatisfactory electrodes will sometimes give accurate readings (Table I).

C. PROBLEMS OF SOIL REDOX MEASUREMENTS 1.

Limit of Eh Measurements

In the case of a reversible chemical reaction involving electrons, the potential (Eh) measured by an inert electrode should reflect the relative concentration of chemical species in the soil (Schu¨ring et al., 2000). The methodology of these measurements is simple in theory. In practice, consideration must be given to many field factors involved with establishing measurements in situ, the Pt electrodes themselves, and their connections with the earth and recording apparatus. The accurate assessment of the redox status using electrodes is still a matter of controversy that is mainly created by the lack of redox equilibrium at the electrode as well as among various redox couples present in a given soil. The problems have been described in numerous publications and will not be repeated here in detail (Grenthe et al., 1992; Grundl, 1994; PfeiVer, 2000; Sigg, 2000; Stumm and Morgan, 1981; Whitfield, 1969, 1974). In brief, the most common problems can be summarized as follows: a. Irreversibility. Irreversibility of redox reactions at the Pt surface of the electrode can be caused by any of the following substances that often accumulate on the surface of the Pt: Pt(OH)2, PtS (Hoare, 1962; Rickman et al., 1968; Whitfield, 1969, 1974), organic coatings (Vershinin and Rozanov, 1983), calcium and magnesium carbonates (Rickman et al., 1968), or aluminosilicates (Devitt et al., 1989). When these substances coat the Pt surface, they cause the potential to drift from true values; this is especially true in well‐drained soils (Bartlett, 1981) and sulfide‐enriched materials (Mansfeldt, 2003; Vershinin and Rozanov, 1983). In spite of these deficiencies, some electrodes perform well for up to 5 years in favorable conditions (Austin and Huddleston, 1999). b. Slow Reaction Kinetics. Slow reaction kinetics of potential‐ determining redox couples can prevent a stable potential from developing (von Langen et al., 1997). As Gao et al. (2002) reported, Eh can drop rapidly under ponding conditions, such as those in a rice paddy, but the chemical reactions with release of iron and other materials created in the new environment will continue for a substantial time, although they may not be initially observed. This condition is typical of systems that are well poised (Bartlett, 1981; Brown, 1934).

SOIL REDOX POTENTIAL

21

c. Mixed Potentials. The measured ‘‘steady state’’ potential is given by the redox system with the highest rate of exchange or corresponds to a composite of two or more redox processes (Bo¨ttcher and Strebel, 1988; Bru¨mmer, 1974; Chadwick and Chorover, 2001; Stumm, 1984), resulting in the restriction of thermodynamic calculations (Frevert, 1984; Machan and Ott, 1972; Rickman et al., 1968).

2.

Technical Problems of Soil Redox Measurements

Additionally, measurements of soil redox potentials associated with technical problems are illustrated in Fig. 5A. a. Reference Cell. Reference cells are often overlooked in the design of a redox system. According to Galster as quoted by Frevert (1984), reactions between internal electrolyte and components of the soil solution can create deviations of up to 60 mV. Strong reducing conditions may liberate H2S, which then diVuses into the reference cell, destroying it by coating the cell with AgS and hence plugging the diaphragm (Cammann, 1977). Reference cells with jelly electrolytes, which represent the current state of technology, salt bridges (Veneman and Pickering, 1983), or reference beddings into microbial resistant material in combination with electrolyte reservoirs (e.g., clinical infusion bags) may overcome these problems (Fiedler and Fischer, 1994). The small size of the electrodes as mentioned above may cause a problem with the galvanic contact. The suggested construction avoids these problems (Fig. 3). Fiedler (1994) reported freezing of the electrolyte, which was solved by using a mixture of electrolyte and glycerin. Such solutions do not change the standard potential of the reference cell, but they do alter the temperature coeYcient favorably. b. Working Electrode. Theoretical problems have been described in numerous publications, but technical problems that might occur with the equipment are not usually published. The life span of electrodes depends on the condition of use, for example water saturation, pH conditions, and mineral chemistry. Because the epoxy allows moisture to interfere with the connection between the wires, junction potential, as mentioned previously, is always a potential problem. Absorptive contamination of the surface of the Pt electrode aVects the measured potential (Vershinin and Rozanov, 1983). It is diYcult to give precise conditions under which the contamination of electrodes might be prevented because the suggestions of published research are contradictory. Cogger et al. (1992) stated that properly constructed and permanently installed soil electrodes were reliable for at least 1 year. Rickman et al. (1968) suggested that adverse eVects due to contamination can be minimized by removing electrodes after

S. FIEDLER ET AL.

22

Data logger; pH/mV meter, and so on Lightning, high voltage static discharges; vandalism

A Transmission line Noise, shorting, corrosion, poor connections, rodents

Amplifier

Corrosion

Pt Reference Soil body studied Cell-body fouled (e.g., pts); little galvanic contact

Contamination; plugging; little galvanic contact

B

Figure 5 DiVerences between permanently installed Pt electrodes (18 months) and noninstalled paired electrodes as aVected by cleaning and concentration of buVer. (A) Experimental and (B) graphical.

SOIL REDOX POTENTIAL

23

2 months in soil and cleaning the Pt sensor. Bailey and Beauchamp (1971) reported a slow response of electrodes during a 6‐month experiment due to sulfide, hydrogen, or organic carbon sorption on the surface of Pt. Whisler et al. (1974) reported that PtS would push the redox potential to more negative values and would result in slow electrode response. Recovered permanently installed electrodes can be tested the same way as newly constructed electrodes. Mansfeldt (1993) found a negligible diVerence between noninstalled electrodes and electrodes that were installed permanently for over 20 months. Austin and Huddleston (1999) noted that electrode testing in quinhydrone solution after removal from the soil showed that 97% of the electrodes still performed within acceptable limits of accuracy after either 3 or 5 years of continuous use. Fiedler (1997) tested electrodes that were installed over an 18‐month period by using diVerent concentrations of test solution (hexacyanoferrate) and found an increasing discrepancy between preinstallation and postremoval testing of electrodes up to 36 mV with a decreasing concentration in paired electrodes (Fig. 5B). These data support statements by Teasdale et al. (1998), who reported that a highly concentrated redox solution is much less sensitive to diVerences between two given electrodes than is the low‐poise reference solution. Whisler et al. (1974) speculated that longevity of Pt electrodes might be inferred by the correspondence between the amplitude of oscillations of electrode readings and the intermittent flooding and drying of soil. Seybold et al. (2002) and Mansfeldt (2003) stated that if properly constructed and permanently installed Pt electrodes are used, soil Eh can be monitored for extended periods under saturated and reducing soil conditions. InsuYcient galvanic contact and polarization of electrodes with the surrounding soils as reported by Yu (1991) is a problem that all researchers might encounter. Cater and Silver (1961) noted that the use of an amplifier with an input impedance <108 O can polarize electrodes so that they record only half their true equilibrium potential. In contrast, a circuit >108 O was only able to polarize the electrodes by a few millivolts. c. Transmission Lines. Plastic coatings on cables are subject to rapid deterioration by solar radiation and general field conditions such as rodents. The deterioration results from adsorption of moisture and shorting of the current. The reaction of ambient moisture reacting with inferior quality plugs and junctions creates corrosion. Corroded transmission lines and connections will result in false signal transmission. The direct soldering of the transmission cables to the data handling boards complicates the maintenance of the apparatus. Therefore, it is recommended that clips, screws, or plug junctions be used instead. Sealing with silicone resins provides excellent insulation. Note that resins containing acetate can increase corrosion. The cables transmitting data to the data loggers are often greater than 10 m and subject to interference (Frevert, 1984). We recommend that the

24

S. FIEDLER ET AL.

electrode cables be shielded against electrostatic coupling such as human and animal bodies, traYc (cars, railroads), atmospheric loads, and other emitters. It is not the relatively low voltages but the low current that makes the signals suVer from electrostatic interferences. Such currents are randomly generated and complicate the interpretation of the signals. At higher amperages, which result from lower impedance, the signals gain immunity. Problems such as interferences due to static, capacitive, or inductive coupling of stray fields can be avoided by using an ‘‘active’’ redox electrode, including an amplifier directly incorporated into the electrode body (Fiedler et al., 2003). This setup eliminates the most important sources of errors that often occur with traditional electrode designs by decoupling the sensitive high‐ impedance primary circuit from all other devices of the signal chain. Similar construction processes were applied for the construction of ion‐selective electrodes (Pe´rez‐Olmos et al., 2001). d. Data‐Recording Unit. Commercial or self‐made multichannel data loggers are commonly used for recording Eh data. However, these units need to be protected against moisture. In particular, it must be ensured that the printed circuit board (PCB) is equipped with an eVective insulation such as spray lacquer (e.g., acetate‐free silicone resin). Sometimes, PCBs that are in contact with high‐impedance primary signals from a working electrode are regarded as the best (e.g., Teflon within ion‐sensitive electrode equipment). However, in the case of Eh, the standard epoxy PCB (R < 1011 O cm1) is suitable for use with circuits processing soil redox potentials. Lightning, static high‐voltage discharges, and voltage spikes are events that may cause permanent damage to electronic devices, especially in complementary metal oxide semiconductor devices (CMOS).

D. TESTING ELECTRODES PRIOR TO INSTALLATION All electrodes need to be tested for accuracy prior to being used in the field. The testing procedure is necessary to ensure that the electrodes are functioning properly and also to place electrodes into groups that will provide the same range in Eh when placed in the soil in multiple plots. We recommend that testing be done in two phases, with each phase using a diVerent test solution. The first phase uses a solution of known and stable redox potential. The ferrous–ferric buVer solution described by Light (1972) will be used for illustration, but other solutions are available (Owens et al., 2005; Tables I and II). Before the testing begins, the Pt wire must be cleaned to remove coatings that will impede electron flow. Coatings can be oxides, organic greases, or numerous other problem materials. The coatings frequently cannot be seen with the

SOIL REDOX POTENTIAL

25

naked eye. To remove coatings, all electrode tips should be rubbed hard or scratched with steel wool. Once cleaned, the electrode tips are placed in a beaker containing the buVer solution along with a reference electrode. Redox potential is then measured for each electrode using the voltmeter. If the solution by Light (1972) is used, then the redox potential reading taken directly from the voltmeter will be þ476 mV. Electrodes falling between 496 and 456 mV are acceptable. Other ranges of acceptability can be defined, depending on the objectives of the work. Electrodes falling outside the acceptable range should not be used. In some cases, failure is due to leaks near the electrode tip, which allow the test solution to contact metal in the electrode other than Pt. The electrodes can occasionally be repaired by resealing the area near the electrode tip with epoxy or heat‐shrink tubing (Owens et al., 2005). Before repairs are attempted, it is helpful to examine the seal around the Pt wire under a microscope so that cracks and holes in the insulation can be identified. The ferrous–ferric buVer solution is well poised, but soil solutions generally are not. Electrodes that appear acceptable in buVer solutions may not perform adequately when placed in the soil. Another test is recommended with tap water before electrodes are selected for field use. Examples of test results for 15 electrodes (five sets in triplicate) are shown in Table III. Redox potential was measured both in the Light (1972) solution and in the tap water. Means and ranges are shown at the bottom of the table. The Light (1972) solution produced data with a narrow range of only 38 mV, indicating that all electrodes would be acceptable for use. Two tests were also done with tap water that was flowing out of an open faucet. Results between the two water tests were similar, but the range was approximately five times as large as that found for the buVered solution. While the results with tap water would seem too variable for the electrodes to be considered reliable, it must be remembered that the tap water is poorly buVered or poised. As a result, the redox potential in the water will change during the testing process.

III. DATA INTERPRETATION A. CORRECTION 1.

OF

FIELD DATA

Correction of Field Data to a Standard Hydrogen Electrode

Generally, redox field data are corrected in order to be relative to the potential of a standard hydrogen electrode by adding a temperature and reference electrode correction to the experimental readings. The temperature correction for various solutions for reference electrodes is given in Table IV. Using an Ag/AgCl reference electrode, for example, the correction factor

S. FIEDLER ET AL.

26

Table III Results of Testing 15 Electrodes in a Well‐Poised Redox BuVer Solution, as well as in Flowing Tap Water Redox potential (uncorrected) (mV) Tap water BuVer solution

Test 1a

Test 2

A A A

481 465 482

360 360 367

363 364 371

B B B

478 476 480

370 375 383

373 386 387

C C C

482 464 482

390 410 410

387 413 413

D D D

481 482 476

420 420 440

413 431 443

E E E

482 482 475

441 457 461

445 461 465

Mean High Low Range

478 482 464 22

377 461 360 101

409 465 363 102

Electrode group

a

This test was used to sort results from lowest readings to highest reading. The tap water is not well poised, but may be closer to soil solution than a redox buVer. The Eh measurements in water are much more variable than with the buVer solution, but can be used to group electrodes into categories that would respond similarly in soil. In this example, three plots having five electrodes each were to be instrumented. Each plot would have one electrode from group A, one from group B, and so on. In this way, the redox potential in each plot would be measured with electrodes responding with the same range in poorly poised solutions.

would be 200 mV for most regions. Note that a 4‐M solution is close to saturation, and readings should not be changed appreciably if saturated solutions (4.5 M) are used in the reference electrodes. 2.

Correction of Field Data for pH

An inert electrode will not only show a positive potential change in response to a change of Eh, but it also responds inversely to changes in pH; as pH increases, Eh decreases. The value most commonly used by

SOIL REDOX POTENTIAL

27

Table IV Standard Half‐Cell Potentials of Selected Reference Electrodes as a Function of Temperature and KCl Reference Solution Concentration Expressed in Moles, Potentials in Millivolts (After Nordstrom and Wilde, 1998) Ag/AgCl Temperature ( C) 10 20 30 40

Calomel

3‐M KCl

3.5‐M KCl

4‐M KCl

3‐M KCl

3.5‐M KCl

4‐M KCl

220 213 205 198

215 208 201 193

214 204 194 184

260 257 253 249

256 252 248 239

254 248 242 234

researchers is 59 mV per pH unit at 25 C for conversion of field observations, which is in accordance with the Nernst‐equation. However, the Eh/pH slope varies within diVerent chemical systems. Experimental values have been found to range from 6 to 256 mV per pH unit. Charoenchamratcheep et al. (1987), for instance, observed that in acid sulfate soil solutions Eh/pH unit slopes between 206 and 256 mV per pH unit occurred. Field mineralogy and solution geochemistry aVect poise and Eh–pH relationships. Therefore, Eh corrections made on the basis of 59 mV/pH unit without synchronous Eh and pH measurements may be unreliable (Collins and Buol, 1970; Garrels and Christ, 1965, p. 176).

B. VARIABILITY Apart from electrode and pH considerations, the interpretation of redox potentials (Eh) proves diYcult due to their variability. Redox potentials can vary spatially and temporally by orders of magnitude, depending on such conditions as changing water table, temperature, and the availability of organic matter for anaerobic microbes. The variability should be recorded in order to provide a basis for the evaluation of the whole redox dynamic in soils. The variability has often been attributed to the imprecision of parallel measurements (Parker et al., 1985). The wide range of Eh potentials (including both natural and artificial changes of Eh) may also create diYculties in the interpretation of the obtained data. 1.

Spatial Variability

The redox potential in soils can vary significantly within (e.g., microsites, Parkin, 1987; Zausig et al., 1993) and between horizons (intrapedonal) (e.g., macro pores caused by earthworm activities, epi and endo saturation

28

S. FIEDLER ET AL.

conditions) as well as between pedons (interpedonal) and can lead to distinct patterns across landscapes. Norrstro¨m (1994) reported that in soil volumes as small as 100 cm3, diVerences between minimum and maximum values of the redox potentials can range between 400 and 1100 mV. Fiedler (1997) states that in general, redox potentials of individual electrodes must be regarded as single‐point or local measurement reflecting the separated, temporally variable compartments that are representative for the surrounding soil at a scale of 1 mm3. Therefore, it is not surprising that Eh can vary up to 800 mV between two electrodes installed in a horizontal distance of 1 cm (Fiedler, 1997). To circumvent the diYculties of extreme spatial variability, a suYciently large number of measurements must be made across a horizon (or depth) in order to account for the natural range of spatial variability (Vepraskas and Faulkner, 2001). However, in many if not most studies, redox probes were applied in duplicate only (Reuter and Bell, 2001; Thompson and Bell, 1996, 1998). Patrick et al. (1996) also recommended using electrodes in duplicate but preferred the installation of electrodes in triplicate. Cogger et al. (1992) recommended the use of 8–10 replicates. Vepraskas and Faulkner (2001) suggested that 10 electrodes per depth are needed to ascertain the precision of the measurement over a complete wetting/drying cycle for soils aVected by fluctuating water tables. The National Technical Committee for Hydric Soils (NTCHS) recommends in their ‘‘Hydric Soil Technical Standard’’ that five replicates be used. Fiedler (2000) found that the diVerences between the redox probes of individual horizons decrease with increasing depth. For example, two electrodes showed a similar trend and level of Eh values (DEhmin ¼ 10 mV, DEhmax ¼ 100 mV) in permanently water‐saturated horizons. Therefore, the use of 10 electrodes in these soil horizons does not seem to be necessary for accurate Eh measurements. In contrast, a higher number of redox probes are needed in horizons that are characterized by highly fluctuating water tables that separate the horizon into a saturated and an unsaturated zone. No clear agreement on the number of required electrodes was found in the literature. In addition, some researchers preferred the installation of electrodes at definite depths (e.g., 25, 50, 75, and 100 cm) (Reuter and Bell, 2003), whereas Blume (1968a,b,c) suggested the installation of the electrodes at diVerent depths across the individual horizons. Despite high spatial variability within individual horizons, similar redox potentials can be observed in certain types of soils. The Eh measurement in diVerent soil types showed that, in general: (1) levels and amplitudes of intrapedonal redox potentials decrease with depth and (2) distinct interpedonal Eh gradients exist between diVerent types of soil. In other words, the intrahorizon Eh diVerences are usually smaller than the interhorizon/soil Eh diVerences. Eh measurements are a used as a tool for: (1) the diVerentiation

SOIL REDOX POTENTIAL

29

of soils and (2) the estimation of dominant redox processes in a certain type of soil. Although water table levels and soil temperature influences are not discussed here, these factors should be included in the data‐gathering process because they also aVect the interpretation of data.

2.

Temporal Variability

The measurement period in published studies has varied from hours (De Mars and Wassen, 1999), a few days (Clay et al., 1990), a few weeks, a few months (Teichert et al., 2000), 1 year (Comerford et al., 1996), 2 years, to as much as 5 years (Reuter and Bell, 2001). These hourly, daily, and weekly measurement intervals depended on the objectives of the diVerent investigations. As noted by Fiedler et al. (2004b), Eh measurements in soils that are aVected by permafrost are only meaningful during (usually short) warm summer periods since freezing can lead to technical problems (e.g., galvanic contact). The investigation of the application of organic fertilizers, which triggers the release of N2O, requires a narrow measurement interval (hours) but a relatively small measurement period (1 month) (Flessa and Beese, 2000). According to USDA‐NRCS (1991), the delineation of hydric soils in the United States is based on the percentage of time with Eh <200 mV during the growth season. The temporal variability of Eh trends varies in relation to the water table, oxygen diVusion (Zausig and Horn, 1992), temperature (Clay et al., 1990), precipitation (Fiedler, 2000), matric potentials (Fiedler, 1997; Karathansis et al., 2003; Thompson and Bell, 1998), and evapotranspiration (Mansfeldt, 2003). Soil redox conditions may be subject to changes according to the following time scales: Diurnal changes can be explained on the basis of the Nernst equation and the van’t HoV’s law (a temperature increase of 10 K enhances the rate of biological reactions by a factor of 2–3). Generally, a temperature maximum is followed by an Eh minimum (Clay et al., 1990; Farrell et al., 1991) with amplitudes of 30–50 mV and a phase shift of 6 h. Event changes can be related to drying/rewetting cycles (McKenzie et al., 1960). The temperature‐related regular variation of the redox potentials may often be concealed by a drastic increase of the water table after precipitation (event changes). In these cases, the Eh amplitudes can range up to 900 mV (Fiedler, 2000). Seasonal changes during summer and winter can be related to gradients in soil temperature, for example depression of microbial respiration at <5 C (Megonigal et al., 1993). The seasonal mean of the redox values can be as much as 300 mV (Fiedler, 1997).

S. FIEDLER ET AL.

30

Annual changes can result from annual diVerences in mean precipitation (corresponding to the water table) and/or temperature of the years of observation (Reuter and Bell, 2001). Mansfeldt (2003) has found mean annual diVerences between 40 and 200 mV based on installation depth in a calcareous Typic Endoaquoll marsh soil over a period of 4 years. Short‐term changes due to soil chemistry result from the production of carbon dioxide and release of hydroxyls in ferric iron reduction. As stated by Ponnamperuma et al. (1966), the pH values of flooded alkali and calcareous soils are considerably lower after 2–3 weeks of flooding, and those of acid soils are considerably higher than their aerobic counterparts. The reduced soils tend toward a range of only 7.2–6.7 (Ponnamperuma et al., 1966). The conditions that exist in the field, therefore, allow for interpretations of the soil’s environmental condition (Schulte‐Eburt and Hofmann, 2000). Besides these systematic Eh changes, an additional irregular shift in redox trends has been observed. These equipment and electronic Eh values, termed here ‘‘artificial’’ Eh values, can lead to false conclusions if they remain unnoticed. However, their identification is relatively easy in the case of ‘‘constant’’ environmental conditions, where artificial Eh values are characterized by spontaneous peaks and/or data outside of the natural range of Eh within soils, as illustrated in Fig. 6.

C.

POOLING

OF

LONG‐TERM DATA SETS

The determination of Eh is usually not of interest in terms of the values themselves; rather, the value of Eh determination is in predicting the behavior of certain soils in their field condition in terms of space and time. It has been demonstrated that redox potentials can vary site specifically throughout the year as a result of annual, seasonal, event‐related, and diurnal changes. Such variation temporally is sensitive to the biological processes of interest in most ecological investigations. In these cases, a designation of ‘‘redox zones’’ in which specific ranges in Eh define electron activities (e.g., ferrous iron production or sulphide production) or categorizing of predominant redox conditions into a distinct ‘‘redox status’’ (e.g., moderately or severely reduced status) is useful for the interpretation of data (Fig. 7). Generally speaking, the interpretation of data collected over longer periods of time has more value and use with the application of basic statistics. Mansfeldt (1993) and Norrstro¨m (1994) calculated the horizon‐specific lower, middle, and upper quartile of Eh values. The relative frequency of Eh values, or cumulative probability distribution of Eh values, was used by Szo¨gi and Hudnall (1998) to develop the concept of relating the frequency of diagnostic wet soil conditions to hydric soil conditions. The probability of

SOIL REDOX POTENTIAL 750

0

24

48

A

600

700

650

Eh (mV)

0 800

24

48

31 800

B

0

600

400

400

200

200

0

0

−200

−200

48

96

48

96

C

600

D

0

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E

F

0 600

−400 −400

−800

400 200

−1200

−800

0

−1600

0

−200

−1200 48 0

−2000 24

48

24

0

Time (h) Figure 6 Examples of natural (regular) Eh potentials, which can be interpreted in relationship to abiotic factors (A–C) and artificial (irregular) Eh values (D–F) that are caused by technical problems: (A) diurnal Eh potential triggered by temperature, (B) Eh decrease after rain, (C) Eh change caused by a fluctuating water table; (D and E) Eh data outside the range of the buVering eVect of water on redox reactions (1 V þ . . . þ 1 V) may result from deflectable transmission line, (F) spontaneous Eh change caused by electro‐smog. ground, and when the soil becomes reduced, the Fe is dissolved oV the pipe (B).

Redox potential (mV) +600

Oxidized

O2 Oxidized

+400

+200

Moderately reduced

NO3−Mn4+

NO3−Mn+4

Oxic

Suboxic

Oxic

Fe+3

Weakly reduced

O2

−20 0

0

Suboxic

Reduced Highly reduced

Reddy et al., 2000

SO42− CO2 Moderately reduced

+Fe+3

Highly reduced

SO42− CO2 Sposito, 1989

Anoxic

Sulfic

Zhi-Guang, 1985

Anoxic

Berner, 1981

Figure 7 DiVerent categories of redox environments with important electron acceptors.

S. FIEDLER ET AL.

32

occurrence of a distinct Eh value (at a certain depth) is obtained by computing the ratio of the number of observations that exceeded the specific Eh of the total number of observations. This ratio is the transformation of a time‐ dependent variable into a frequency and can be represented by a cumulative probability function. By focusing on key Eh thresholds, Fiedler and Sommer (2000) established a useful interpretive tool for the estimation of redox dynamics of hydromorphic soils by using the percentage of time the Eh values were above or below the threshold values annually. These researchers demonstrated that their procedure was useful for horizons in pedons and pedons in landscapes. The key thresholds these researchers utilized were Mn‐reduction using Eh < 450 mV; Fe‐reduction using Eh < 170 mV; and CH4‐oxidation using Eh > 75 mV. These key thresholds (e.g., CH4‐oxidation) can be used as input parameters for simulation models, such as wetland‐DNDC, as confirmed by Zhang et al. (2002). In addition to the estimation of element/matter dynamics in soils, Eh thresholds allow the classification of soils. The technical guidance for hydric soil delineations (Hurt and Carlisle, 2001; Hurt et al., 2002; NTCHS, 2004) relies on redox conditions as one of the major criteria for hydric soil identification. The hydrology criterion for wetland delineation should be considered fulfilled when reducing conditions (threshold of Eh7 200 mV) are continuously available (>5% of growing season) in the upper part of the soil (operationally defined as above 30 cm). According to the World Reference Based Soil Classification system (FAO‐WRB, 1998), stagnic properties are defined below a threshold of rH < 19, where rH ¼ (Eh/59) þ 2pH). In Soil Taxonomy (Soil Survey StaV, 1999), reduction and oxidation are both required for ‘‘aquic’’ conditions but the duration of the reduction is not specified.

IV. ALTERNATIVE METHODS FOR ASSESSING REDUCTION IN THE FIELD A.

IRON‐COATED (IRIS) TUBES

Jenkinson and Franzmeier (2006) proposed using iron‐coated tubes to assess the redox potential in soils. They put a thin coating of ferrihydrite (Fe(OH)3) onto the surface of a PVC pipe (Fig. 8A). The coated pipe simulated a rusty steel nail. It was named an IRIS tube for ‘‘indicator of reduction in soils’’ by Jenkinson and Franzmeier (2006). In reduced soil, microbial oxidation of organic matter produces materials that can reduce and dissolve the Fe oV the IRIS tubes (Fig. 8B). The IRIS tubes have an

SOIL REDOX POTENTIAL

33

A Entire tube is ~1.5 in. i.d. (1.27 -cm diameter) PVC pipe (Schedule 40)

15 cm or more above ground 30 cm long zone coated with a single layer of Fe hydroxide "paint"

Pipe diameter is not critical but should be selected to make contact with soil when inserted into hole

B

This zone should be placed below ground

IRIS tube operation White spots on tube show soil was reduced

e−

e−

Fe2+

e−

Fe2+

e−

Fe2+

F e2+ F e2+

H+ Figure 8 (A) The basic parts of an IRIS tube are an above ground area that remains bare and a 30 cm or so long area of the PVC pipe that is coated with an Fe(OH)3 paint. (B) The IRIS tube is inserted into the soil such that all the Fe(OH)3 paint is in the soil.

orange color when manufactured due to the ferrihydrite. When the Fe is reduced, it dissolves oV the tube and the place of Fe loss is seen as a white spot, the color of the underlying PVC pipe. The soil redox potential is estimated as being either low enough to reduce Fe or high enough to keep it in an oxidized form on the tube. The advantages of using IRIS tubes over direct redox potential measurements include being simpler and less expensive to conduct, recording reducing conditions continuously without direct measurement, and showing the location and pattern of reduction. White spots of Fe removal on the tubes show that the reduction occurred in small microsites, as opposed to complete reduction of a horizon whereby all Fe would

34

S. FIEDLER ET AL.

be removed from the tube. The amount of reduction that has occurred is estimated from the area of Fe removed. The IRIS tube is also an excellent tool to use in remote locations where frequent visits are not possible to assess reducing conditions. The IRIS tubes are easy to use. Jenkinson and Franzmeier (2006) augered a hole into the soil where data were needed and inserted the IRIS tube into the hole. While direct contact with the soil is necessary, the contact should not be so tight that Fe is scraped oV the tube as it is inserted or removed for examination. The amount of Fe loss from the tube can be assessed weekly or as often as time permits. To measure the amount of Fe loss, Jenkinson and Franzmeier (2006) photographed the IRIS tubes on all sides, downloaded the pictures to a computer, and used software to estimate the area of the pipe that lost Fe.

B. DYES Solutions of some organic dyes react with redox couples and turn color. Childs (1981) proposed using a 0.1% solution of a,a0 ‐dipyridyl dye to identify when Fe2þ is present. The USDA‐NRCS uses a 0.2% solution of a,a0 ‐ dipyridyl dye in a 1‐M solution of ammonium acetate (Vepraskas, 2002). The solution is clear before use; however, when it is dripped or sprayed onto a freshly extracted sample of soil, it will turn pink if reduced Fe is present. The reaction occurs in 30 s or less. The dye only reacts with Fe2þ and thus is of little use in soils that contain no Fe. This dye has proven a simple and eVective way to evaluate whether reduced Fe is present in soils. It must be used on freshly extracted samples of saturated soils, and the sample must not be allowed to dry before the dye is applied; otherwise, the reduced Fe will oxidize. False positive reactions can occur if the soil to which dye is applied contains metal fragments from steel sampling equipment, such as augers or shovels. Samples containing high levels of organic matter may contain Fe as a ferric–organic complex (Childs, 1981). When the dye is sprayed onto such samples and exposed to direct sunlight, the organic compounds may oxidize and reduce the Fe in the sample. This reaction is capable of causing the dye to turn red even when the freshly extracted material contained no reduced Fe. In another alternative method, which works well with dark soils that can mask or hide the color in the above method and also evaluates ferrous iron attached on the cation‐exchange complex, a solution of ortho‐phenanthroline is used with a small amount of NaCl in a small glass container like a test tube. A small amount of soil is added, and the mixture is shaken. If any ferrous iron is present in that soil, the supernatant solution will be pink (Richardson and Hole, 1979).

SOIL REDOX POTENTIAL

35

C. ZERO VALENCE IRON RODS Starkey and Wight (1946) did not observe any corrosion of iron pipes when the Eh values were less than 400 mV. McKee (1978) proposed using uncoated steel welding rods to measure water table depth. Iron forms rust in the oxidized (Fe3þ) conditions above the water table but is reduced under anaerobic conditions to Fe2þ with a lack of rust coloration. McKee (1978) believes that the water table is the boundary between these two conditions and can be recognized by red above and no rust coloration below. Carnell and Anderson (1986) measured rusting on steel rods in a Stagnopodzol. They determined that the maximum rust depth correlated well with both water level and rooting depth of Sitka spruce (Picea sithchensis). Carnell and Anderson (1986) observed that the lower level of brownish coloration on the metal correlated with water table depths. Bridgham et al. (1991) studied the suitability of steel rod rusting as a hydrologic indicator in wetland and nonwetland soils. They used both field and microcosm approaches with Histosols and one mineral soil type (Typic Fluvaquent). The steel rods worked well with a water table that did not fluctuate rapidly. Rapid fluctuations did not allow enough time to reduce and dissolve rust when it had formed. The oxidation of the rod also required considerable time because the drainage was slow and the oxygen was hindered in returning to the pores. They concluded that the steel rod technique was valuable as an indicator of hydrology for basic science studies. They thought the technique was not suitable for jurisdictional wetland determination in zones with fluctuating water tables due to the lag periods in rod oxidation and the time needed for previously formed rust to dissolve when ponded again. Owens et al. (1999) used buried iron metal rods for estimating anaerobic and aerobic conditions in soils. Later, Owens (2001) stated that the idea may have implications in hydric soil determination, horticultural and agricultural applications, and remediation techniques. Owens used zero valence iron rods placed in three soil toposequences. His observations of the surface coatings formed by oxidation were compared to piezometer, soil moisture, a,a0 ‐ dipyridyl reactions, oxygen concentration, and redox potential measurements every 3 weeks for 1 year. He believes that the coatings on iron metal rods inserted in the soil appear to correlate well with specific oxygen ranges. Rods in soils with O2 concentrations below about 2% do not develop bright (7.5YR 4/4 to 5/8) oxide/oxyhydroxide coatings but instead formed black (10YR 2/1 to 2/2) coatings. Rods in soils with O2 concentrations between about 2 and 5% develop variegated bright (7.5YR 4/4 to 5/8) oxide/ oxyhydroxide coatings, indicating microsite diVerences in O2 concentrations. Rods in soils with O2 concentrations above about 5% with adequate moisture are almost completely coated with bright (7.5YR 4/4 to 5/8) iron

S. FIEDLER ET AL.

36

oxide/oxyhydroxides. Once iron oxide/oxyhydroxides were formed on the rod, they were not reduced during subsequent reducing conditions. Owens concludes that these data indicate the iron rods can be used to provide information on a bracketed time period rather than an integrated measurement. Vepraskas et al. (1999) noted that the dissolution of iron was slow and required soils with considerable organic matter available for microbial use. Over a short time period (about 1 week), they observed that the soils having a high content of organic matter had over 10 times the ferrous iron, whereas soils having a low content of organic matter had only a trace amount of ferrous iron. Formation of rust is endothermic and easier to achieve than the reverse.

V.

FIELD INSTALLATION AND PROCEDURES FOR REDOX POTENTIAL MEASUREMENTS

Redox potential measurements are made with three pieces of equipment: Pt electrodes, a reference electrode, and a device to read voltage, such as a voltmeter (Fig. 2). The basic components of the two electrodes are shown in Figs. 2 and 3. Pt electrodes can be purchased or made by hand. Methods for constructing electrodes have been described previously in this text and in Faulkner et al. (1989), Mueller et al. (1985), Patrick et al. (1996), and Vepraskas (2002). Reference electrodes are usually purchased, but the construction of a stable field type (Ag/AgCl type) is described in Section II. Because the electrodes will be in contact with the soil, they should be rugged enough for field use, which is the case with the Fiedler design described in Section II (Fiedler et al., 2003). The voltmeter must be capable of measuring millivolts. Laboratory grade Eh/pH meters are frequently used, but commercial voltmeters also work. A field installation of the Pt electrode, reference electrode, and voltmeter is shown in Fig. 2. The Pt electrodes may be installed and left in place for up to a year, with notable exceptions (Section III). The reference electrode and voltmeter are taken to the field for each measurement.

A. NUMBER OF ELECTRODES

TO

INSTALL

Redox potential is an extremely variable property. Figure 9 shows the range in Eh that was found in one soil that was saturated for a portion of the year in a horizon that contained five Pt electrodes. The graph shows the mean and range in redox potential over time. Before the soil horizon that contained the redox Pt electrodes became saturated, the Eh was in a fairly

SOIL REDOX POTENTIAL

37

Figure 9 Variation in redox potential for a soil that undergoes periods of saturation and drainage. Means of five electrodes are shown for a mineral soil at a depth of 30 cm. The largest range in redox potential occurs soon after soil saturates or drains. Modified from Vepraskas and Faulkner, 2001.

narrow range. The variability in redox potential increased when the water table rose and redox potential decreased. The high variability was a result of reducing conditions that occurred in the soil around some electrodes but not in others. Possible causes for high variability in redox potential are illustrated in Fig. 10. When soils are saturated quickly, some air is trapped in bubbles in the soil. As shown in Fig. 10, any electrodes that have their tips in bubbles that contain oxygen will record high redox potentials. On the other hand, electrodes that are placed near pieces of organic C that are being oxidized by bacteria will be in reduced soil and will record low redox potentials. Intermediate redox potentials will occur when electrodes are placed somewhere in between these extreme conditions. If saturation is maintained for a long enough period, then the range in Eh values narrows as the entire horizon becomes reduced. See both Sections II and III for comments on other technical problems. The broad range in Eh values is typical and shows that to measure the Eh of a soil horizon, an adequate number of electrodes must be installed. Statistical techniques could be used to determine a recommended number of electrodes, but a general rule of thumb is to install between 5 and 10

S. FIEDLER ET AL.

38 Low Eh

High Eh Intermediate EH

Fe2+

O2

O2

Saturated soil

Figure 10 Illustration of possible explanations for why Eh measurements can be variable across lateral distances of 1 m or less. The soil is saturated, but air bubbles containing oxygen gas persist because the water table rose quickly and trapped air. Platinum electrodes inserted into the trapped air bubble may record high Eh values. Near a root, bacteria may be oxidizing organic C and releasing electrons that are used to reduce Fe. An electrode inserted next to the root may record a low Eh value. Other electrodes that are not near a root or gas bubble may record intermediate Eh values.

electrodes for each depth of interest because of the variation explained earlier and given in Fig. 10. Using fewer than five electrodes will probably not account for all of the variability, and using more than 10 can be expensive. In addition, researchers must be aware that some electrodes will fail, and so replicated electrodes are also necessary to ensure that suYcient data are recorded when electrode failures occur. All electrodes need to be thoroughly tested before installation (Section II.D).

B. INSTALLING Pt ELECTRODES The depth of Pt electrode installation will depend on the objectives of the research. Studies on wetlands typically focus on the upper 30 cm of the soil, whereas studies related to soil genesis may focus on a greater range of depths. For any application, electrodes should be installed in ways that will prevent oxygen penetrating from the surface to the electrode tip and also so that the Pt wire is in contact with natural soil. Pt electrodes can be installed in three ways to make readings: (1) direct insertion, (2) rigid rod, and (3) slurry‐sealed. Direct insertion works best in soft materials that are

SOIL REDOX POTENTIAL

39

under water. The electrode is pushed into the soil, and a reading is taken with the voltmeter and reference electrode. Being under water, there is little chance that air will enter the soil as the electrode is being pushed into the soil. In harder materials, some workers push a rigid rod into the soil to make a hole slightly smaller than the electrode. When the rod is removed from the soil, the Pt electrode is slid into the hole. A tight fit is believed to seal against oxygen penetration. This method probably works best when shallow readings are made. The third way to install electrodes is shown in Fig. 11. A hole is first bored to within 1 cm of the desired depth of installation. The hole can be 2 cm or more in diameter. Soil at the depth of the electrode tip is then used to make slurry of mud with water. The slurry is poured into the hole and the electrode pushed through the slurry into natural soil. To complete the seal, more slurry is added to the hole until the hole is filled around the electrode. It is important that the slurry be taken from onsite and from the depth of the electrode tips to ensure that the soil chemistry (particularly the pH and organic C content) is not altered by the slurry.

Auger hole to within 1 cm of desired depth of electrode

Fill with slurry Insert electrode, push tip into natural soil Figure 11 Installation of a Pt electrode. After the hole is bored into the soil, it is filled with a slurry made from the extracted soil. Only soil from the horizon that will contain the Pt tip should be used. Once the hole is filled with slurry, the Pt electrode is pushed through the slurry and the Pt wire seated in natural soil.

40

S. FIEDLER ET AL.

The basic process used to install five electrodes at a depth of 25 cm is as follows: 1. Auger five holes to a depth of 24 cm using a push‐probe. Holes should be about 30 cm apart and are generally put in a line. 2. Fill each hole with a mud slurry made from the soil removed. Do not use the subsoil material below 25 cm because it will diVer chemically from the topsoil. 3. Insert one electrode into a filled hole and push to the bottom of the hole. Try to seat the Pt into natural soil at the bottom of the hole, but do not push so hard that you bend the wire. 4. Pour more slurry around the wire to keep the hole filled. 5. Make one more hole to a depth of 12 cm, but do not put slurry into it. This hole will be for the reference electrode. It should be placed about halfway along the line of Pt electrodes, about 30 cm from the middle electrode. The distance between electrodes that are installed at the same depth is determined for convenience. The electrodes should be close enough so that they can be read without moving the reference electrode, yet not so close that one electrode is disturbed when an adjacent electrode is installed. Before they are inserted into the soil, the electrodes can be rechecked for accuracy by measuring the voltage of a buVer solution. This step ensures that electrodes were not damaged in transit to the field. To test the equipment: 1. 2. 3. 4.

Brush the Pt wires with steel wool to clean them. Insert the clean electrodes into the test solution. Insert the reference electrode (remove black cap covering ceramic tip). Connect black wire to ‘‘common’’ terminal on voltmeter and attach to reference electrode. 5. Connect red wire to ‘‘volt’’ terminal on voltmeter and connect to a Pt electrode. 6. Turn meter on, and read voltage. Record voltage. 7. If voltage is within 20 mV of the number on the test solution bottle, the electrodes and the rest of the equipment are good. Once installed, electrodes may be left in place for up to a year. Each electrode should be checked for accuracy in a buVer solution each year to ensure that the electrodes are functioning properly. It is advisable to secure the electrodes to racks made of PVC pipe to keep the electrodes from being stepped on and moved. Wires may also be attached to each electrode with the other end of the wire secured to a wooden table. This arrangement prevents workers from compacting the soil near the installation and also lessens the need for bending over.

SOIL REDOX POTENTIAL

C. READING

THE

41

ELECTRODES

Electrodes installed by direct insertion can be read as soon as the electrode tip reaches the proper depth. When electrodes are installed by using a rigid rod or by mud slurry, then readings should be delayed at least 24 h to allow the solution around the Pt wire to equilibrate with the soil solution. The basic procedure for measuring redox potential is: 1. Place some water in the open hole, dug to 10 cm, and insert the reference electrode. Be sure you took oV the protective plastic caps that cover the base of the electrode. 2. Connect the black wire from the voltmeter (common terminal) to the reference electrode. 3. Connect the red wire from the voltmeter to one Pt electrode. 4. Turn the meter on and wait for the readings to stabilize. 5. Record the voltage. Note whether the voltage is ‘‘plus’’ or ‘‘minus.’’ 6. Convert the voltage to ‘‘Eh’’ or redox potential by adding to the voltage reading a correction factor that is appropriate to the reference electrode. For most work, a correction factor of þ200 mV can be used for Ag/AgCl electrodes, þ250 mV for calomel electrodes.

D. COMMON FIELD PROBLEMS The most common problem that occurs while making redox potential measurements is that the voltage does not stabilize but steadily moves, or drifts, to higher or lower values. The drift can be caused by natural processes that include a gradual change in the chemistry of the soil solution around the electrode tip. Drift may also be caused by human error. When it does occur, workers should check that the voltmeter is connected properly to the reference electrode and Pt electrode. The contact between the reference electrode and soil should also be checked. The voltage can drift if the circuit is broken among voltmeter, reference electrode, and Pt electrode. In many cases, the voltage will drift when the kinds and amounts of oxidized and reduced species around the Pt tip are changing. This has to be expected whenever the soil water content changes from unsaturated to saturated or the reverse. The greatest amount of drift may be expected when the water table fluctuates and the soil horizon either saturates (and air bubbles may become trapped) or drains and oxygen penetrates along large cracks and root channels while the matrix is reduced. Redox potentials at a given electrode are stable when oxygen is the dominant electron acceptor, as is commonly the case in unsaturated soils. When saturation occurs, bacteria that are respiring near the tip of the Pt electrode will deplete the soil water of oxygen and eventually other

42

S. FIEDLER ET AL.

electron acceptors will be used. As such substances as nitrate, manganese oxides, and iron oxides are reduced, the redox potential will fall and may cause drift. If soil water is moving past the Pt tip, the composition of the soil solution near the tip will slowly change, which can also cause drift. Voltage drift is a frustrating problem to resolve in the field, and workers are advised to use a standard procedure when making measurements. At least two options are available to address drift. The first is to leave the voltmeter connected to an electrode until the drift stops, and then record the redox potential. This may require 15 min or longer. When many electrodes must be read, such a long waiting period may not be feasible; however, if it is critical to know the exact redox potential on a given day, then this approach may have to be used. An alternative procedure is to allow a set time, for example, 1 min, for each electrode to stabilize and then take the reading. This approach is reasonable when changes in redox potential are being monitored seasonally or over long time periods. Drift that occurs on a given day may simply be part of a long‐term trend as the soil is changing from oxidized to a reduced condition. Such trends become apparent if data are collected over the course of months.

E. FIELD PH MEASUREMENTS A pH value is needed to interpret redox potentials. Reducing reactions tend to convert Hþ to water as shown by the half‐reaction for the reduction of Fe: Fe2 O3 þ e þ Hþ ! Fe2þ þ H2 O When soils become saturated, the soil pH shifts toward 7 in both acid and basic soils (Ponnamperuma, 1972). In acid soils, the amount of pH change is related to the amount of Fe reduction that occurs, while in basic soils the pH shift toward 7 is controlled by the accumulation of CO2 gas. In many field soils, the pH may change by less than 2 pH units after several weeks of saturation and reduction. However, the pH change can be more in rice paddies where large amounts of organic matter and iron oxides allow for large amounts of Fe oxides to be reduced (Ponnamperuma, 1972). When soils drain, the pH shifts again toward its original value as Fe2þ is oxidized or the concentration of CO2 gas is lowered. Soil pH should be measured in the field on samples of either soil or water. Allowing samples to oxidize or dissolved gases to escape can cause pH to change when exposed to air. Measuring pH electrometrically is most accurate. Water samples may be extracted from piezometers placed at the same depth as the electrodes. When water samples are used, the piezometer should be emptied and allowed to refill before a sample is removed for pH

SOIL REDOX POTENTIAL

43

measurements. For soil samples, pH should be measured on soil obtained from the same depth as the electrode tip. Soil pH is commonly measured on a slurry of 1:1 soil to water by volume.

F. INTERPRETING REDOX POTENTIAL Redox data are interpreted using an Eh–pH diagram that relates a given reducing half‐reaction to redox potential. An example is shown in Fig. 12 for the reduction of Fe(OH)3. When the soil pH is known, then the redox potential from an electrode can be plotted on the diagram to determine the form of Fe that should exist in the soil. Diagrams such as that shown in Fig. 12 can be prepared for any half‐reaction of interest (McBride, 1994; Vepraskas and Faulkner, 2001). Field experience has shown that such diagrams work well for Fe minerals. The theoretical redox potential at which Fe reduction occurs, shown on the Eh–pH diagram, is close to the actual redox potential where reduction has been observed (McBride, 1994). Similar results have been found for the reduction of SO4 and CO2. On the other hand, theoretical redox potentials for the reduction of O2, NO3, and MnO4 have been found to be markedly higher than the redox potentials found in nature. Accordingly, when redox potentials are measured to assess the reduction of O2, NO3, and MnO4, it may be advisable to use a phase diagram such as that for Fe, and then assume that if Fe has been reduced, then O2, NO3, and MnO4 must have been reduced as well.

Theoretical Eh values in literature

1200

Aerobic Eh (mV)

800

O2

H2 O

400

0

Fe

(O

Fe reduced

3

5

Fe 2

pH

7

H)

+

Anaerobic, Fe oxidized

3

9

Figure 12 An Eh–pH diagram that can be used to interpret redox potentials. The soil pH must be known before the field‐measured Eh values can be interpreted. This graph can be used to identify the critical Eh values at which Fe reduction occurs in a soil of a given pH. Such data are then used to determine when Fe reduction occurred in a soil (McBride, 1994).

44

S. FIEDLER ET AL.

VI. SUMMARY Redox potential measurements (Eh) are the tendency of a pair of chemical compounds to undergo a transfer of electrons. In hydric soils, the pair of ions of most interest is Fe2þ/Fe3þ. Redox measurements have never been more important than at present because of the need for better measures for hydric soils. The various levels of reduction and oxidation relate to important environmental conditions. The interpretation of redox potentials obtained from redox potential measurements in soils is subject to many problems (e.g., chemical restraints and technical problems). It is diYcult if not impossible to interpret redox reactions in soils in a quantitative way by applying basic equilibrium thermodynamics. However, this does not preclude the usefulness of these estimates for understanding how the ecosystem functions. Therefore, the attempt to identify dominant redox processes in a given soil is of scientific interest as well as practical relevance. Many practical, persistent, and controversial problems exist. One problem is that the iron ion pair is not the only pair that may undergo an electron transfer in field conditions. Commercial working and reference electrodes are often not suitable or rugged enough for field use in soils. Electrode construction, insulation, and contamination are also problems. Suggestions were given on electrode construction, which should help with these problems. Insulation with epoxy products, which often adsorb water, has been unsatisfactory. We suggest alternatives. Cleaning to prevent contamination and standardization procedures of working electrodes are suggested. Standard field installation procedures, such as those suggested here, are necessary to ensure consistent results. Additionally, field pH, water table, and soil temperature data are needed for interpretation of the results. Numerous repetitions of each soil element measured are needed because of these variations, and these repetitions are costly. Currently, in situ Eh measurements are still the most eYcient way of estimating reducing reactions for virtually any soil. In the future, evaluations of redox measurements should include comparisons to ferrous iron dyes, such as dipyridyl, comparison to ‘‘IRIS’’ or ‘‘indicator of reduction in soils,’’ and to zero valence iron materials. Such comparisons with field pH measurements should lead to better interpretations of soil conditions.

REFERENCES Abdelouas, A., Lutze, W., and Nuttall, E. H. (1998). Chemical reaction of uranium in groundwater at a mill tailing site. J. Contam. Hydrol. 34, 343–361. Abdelouas, A., Lutze, W., Gong, W., Nuttall, E. H., Strrietelmeier, B. A., and Travis, B. J. (2000). Biological reduction of uranium in groundwater and surface soil. Sci. Total Environ. 250, 21–35.

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INCREASING RICE PRODUCTION IN SUB‐SAHARAN AFRICA: CHALLENGES AND OPPORTUNITIES V. Balasubramanian,1, M. Sie,2 R. J. Hijmans1 and K. Otsuka3,4 1

International Rice Research Institute (IRRI), Metro Manila, Philippines Africa Rice Center (WARDA), 01 BP 2031 Cotonou, Benin, West Africa 3 Foundation for Advanced Studies on International Development (FASID), 7‐22‐1 Roppongi, Minatoku, Tokyo 106‐8677, Japan 4 National Graduate Institute for Policy Studies, 7‐22‐1 Roppongi, Minatoku, Tokyo 106‐8677, Japan 2

I. Introduction II. Rice Demand and Supply III. Wetlands: The Potential Resource for Rice Production in SSA A. Definition, Area, and Distribution of Wetlands B. Types and Characteristics of Wetlands IV. Rice Soil Resources A. Dryland Soils and Their Characteristics B. Wetland Soils and Their Characteristics V. Agroclimatic Zones and Rice Ecosystems A. Dryland Rice Ecosystems B. Wetland Rice Ecosystems VI. Rice Production Constraints in SSA A. Physical, Biological, and Management Constraints B. Human Resource Constraints C. Socioeconomic and Policy Constraints VII. Rice Research and Technology Development During the Past 20 Years A. Rice Germplasm, Breeding, and Variety Development B. Rice Seed Production and Distribution Services C. Crop Establishment D. Nutrient Management E. Water Management for Rainfed and Irrigated Areas F. Weeds, Insect Pests, and Diseases and Their Management G. Grain Quality Management: From Breeding to Milling

Present Address: Freelance consultant, Ramya Illam, 42 Thadagam Road, Velandipalayam Post, Coimbatore 641025, India.

55 Advances in Agronomy, Volume 94 Copyright 2007, Elsevier Inc. All rights reserved. 0065-2113/07 $35.00 DOI: 10.1016/S0065-2113(06)94002-4

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V. BALASUBRAMANIAN ET AL. H. Diversification of Rice Farming Systems I. ICM for Rice VIII. Rice Intensification Issues and Thoughts for the Future A. Rice Intensification in Relation to Vector‐Borne Human Diseases B. Environmental Issues Related to Rice Intensification in SSA C. Preparing for the Impact of Climate Change D. Technology Delivery and Deployment Issues E. Policy Support for Rice Intensification in SSA IX. Conclusions: Challenges to and Opportunities for Enhancing Rice Production in SSA Acknowledgments References

Sub‐Saharan Africa (SSA) faces multiple problems. The main one is improving the lives of the 30% of its population that suVers from extreme poverty and food insecurity. As more than 70% of the population lives oV farming and related activities, agricultural development will have to play a major role in improving this situation. Fortunately, Africa has an abundant supply of natural resources that can support a huge expansion in food, specifically rice production. Because of strong demand, rice area expansion in SSA is larger than for any other crop. Total milled rice production increased from 2.2 million Mg in 1961 to 9.1 million Mg in 2004. Rice imports into SSA also increased from 0.5 million Mg of milled rice in 1961 to 6.0 million Mg in 2003 and SSA currently accounts for 25% of global rice imports, at a cost of more than US$1.5 billion per year. Therefore, many African governments accord high priority to developing their local rice sector as an important component of national food security, economic growth, and poverty alleviation. The abundant supply of agroclimatically suitable wetlands (239 million ha) and water resources can support a large expansion in rice area and productivity. Currently, less than 5% of the potentially suitable wetlands are planted with rice because of various constraints. Expansion and intensification of rice cultivation in SSA will not compete with other crops in terms of land and water resources because, during the rainy season, only rice can be grown on low‐lying wetlands, including inland valleys. In addition, the labor‐intensive nature of rice cultivation will provide additional sources of work and income to the rural poor, especially women. Should labor shortages become acute, however, appropriate mechanization can be considered. Small farmers want to earn money from rice farming, but lack modern inputs and capital to fully exploit their rice lands as these items are limited or not available. This is where an innovative public–private partnership is desirable to support the intensification of rice farming. Rice is cultivated in four ecosystems of SSA: dryland (38% of the cultivated rice area), rainfed wetland (33%), deepwater and mangrove swamps (9%), and irrigated wetland (20%). Many abiotic stresses (drought, flood, and variable rainfall; extreme temperatures; salinity; acidity/alkalinity and poor soils, soil erosion, and high P fixation) and biotic constraints [weeds, blast, Rice yellow mottle virus (RYMV), and African rice gall midge (AfRGM)] limit rice production on the continent. The changing climate is expected to further aggravate the abiotic constraints and reduce

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rice yields in all ecosystems. Rice production is also restricted by many technical, management, socioeconomic, health, and policy constraints. The constraints to irrigated wetland rice in the Sahel of SSA are similar to those faced by Asian farmers in the 1960s; therefore, well‐tested irrigated rice technologies from Asia and elsewhere are being introduced and adapted to local conditions to obtain fast returns on investment. For rice in irrigated wetlands in the humid and moist savanna zones, rainfed wetlands, and drylands, locally developed NERICA (new rice for Africa) varieties and production technologies are being tested in target environments. The progenies of Oryza glaberrima and O. sativa subspecies indica are better adapted to rainfed and irrigated wetlands, while those of O. glaberrima and O. sativa subspecies japonica are more suited to rainfed drylands. In addition, research is ongoing to tackle SSA‐specific problems such as RYMV and AfRGM and to develop eYcient crop management technologies. Currently available best management practices (integrated crop management options) for diVerent rice ecosystems are shown in Table XV. Additional support through the provision of technical advice through revamped national R&D services; a supply of good‐quality seed and other inputs, including farm credit; and enabling policy are needed for profitable and sustainable intensification of rice cultivation in SSA. It is also critical to organize preventive health measures for farmers against wetland‐related diseases (malaria, bilharzia, and so on), protect certain natural wetlands (e.g., with bird sanctuaries), preserve mangrove forests in strategic coastal belts and rich peats in inlands, and use chemical inputs eYciently to minimize pollution and maintain environmental quality while intensifying rice production. Anticipatory research is needed to tackle the impacts of changing climate on rice farming and the environment. Modern information and communication technologies (ICTs) can be exploited to reach out to farmers in remote areas and to deploy technologies eVectively. In addition, the development of innovative private– public partnerships and the organization of farmers into user‐groups will enhance the training, farmer education, and technology adoption required for intensive commercial rice farming. # 2007, Elsevier Inc.

I. INTRODUCTION Sub‐Saharan Africa (SSA) is the world’s poorest region. More than 30% of the 900 million people living in SSA suVer from pernicious hunger and malnutrition (Farm‐Africa, 2004; IAC, 2005; Rosegrant et al., 2005; Sanchez and Swaminathan, 2005). The prevalence of diseases such as malaria and AIDS is very high (FAO, 2006b) and many countries have been disrupted by civil war. Although the current situation is grim, there is also much potential to improve this situation. More than 70% of the population of SSA is rural, and agricultural development is essential to achieve economic growth, poverty alleviation,

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and food security. The adoption of more productive agricultural practices coupled with the development of rural infrastructure and local markets and supportive agricultural policy is crucial to improving both rural and urban living conditions. Achieving this requires that donor funds and a significant part of national income be strategically invested in scientific agriculture and farming methods and revamping national agricultural research and extension systems (NARES) as a means of improving rural livelihoods and national food security. African agriculture consists of a diverse set of farming systems that have arisen in response to the large variations in ecological, social, and economic conditions. Dixon et al. (2001) delineated 15 broad farming systems, including forest‐based systems; systems dominated by livestock, cereals, and root crops; and mixed systems. It is thus evident that improving the production of several crops and livestock will have a role in agricultural development in SSA. In this chapter, we focus on the role of a single crop, rice, that we think will have a particularly important role in this process. From a low base, rice consumption and production have increased tremendously in SSA over the past decades, and this trend is expected to continue. Moreover, rice can be very productive and sustainable and be produced in areas where other crops cannot be grown. However, unlike in Asia, where the unique and uniform rice‐based farming system benefited from Green Revolution technologies during 1965–2000, productivity gains in African rice farming will come in small increments due to Africa’s diverse nature of cropping systems. Yet the potential for growth in the African rice sector is enormous. In this chapter, we first assess the potential resources for rice production and characterize the rice‐growing environments in SSA. We then discuss the current status of rice farming and related production constraints in four major ecosystems, assess the progress of rice research and development in the past 20 years, and finally indicate the challenges and opportunities for rice intensification in SSA.

II.

RICE DEMAND AND SUPPLY

Rice is a traditional staple food in parts of West Africa and Madagascar, and it is increasingly becoming an important food in East, Central, and Southern Africa. In recent years, the relative growth in demand for rice is faster in SSA than anywhere else in the world. Demand for rice has been growing due to population growth and a shift in consumer preference for rice, especially in urban areas. Annual per capita milled rice consumption in SSA has increased from 11 kg in 1961 to 22 kg in 2003 (Fig. 1). Rice consumption increased steadily in all countries, except Madagascar. Mean per capita

Milled rice consumption (kg per capita)

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200 180 160 140 120 100 80 60 40 20 0 1961

1966

1971

1976

1981

1986

1991

1996

2001

Year Côte d’Ivoire Nigeria Sub-Saharan Africa

Kenya Senegal

Madagascar South Africa

Figure 1 Trends in per capita rice consumption in SSA, 1961–2001 (source: FAO‐STAT, FAO, 2006a).

rice consumption is high in Madagascar (122 kg), Guinea Bissau (103 kg), Coˆte d’Ivoire (Ivory Coast; 100 kg), Senegal (100 kg), Sierra Leone (93 kg), the Gambia (90 kg), Guinea (73 kg), and Gabon (72 kg). Local rice production cannot meet the increasing demand for rice in many African countries (Hossain, 2006). Although milled rice production increased from 2.2 million Mg in 1961 to 8.7 million Mg in 2004, rice imports also increased from 0.5 million Mg in 1961 to 7.4 million Mg in 2004 (Fig. 2) (FAO‐STAT, FAO, 2006a; IRRI, 2006). In 2002, four of the six largest rice importers were Coˆte d’Ivoire, Nigeria, Senegal, and South Africa. Currently, SSA accounts for 25% of global rice imports at a cost of more than US$1.5 billion per year. Projected rice imports into West Africa alone will be between 6.5 and 10.1 million Mg in 2020 (Lanc¸on and Erenstein, 2002). Declining global rice stocks and the predicted doubling of the rice price by 2008 will put additional strains on rice‐importing countries in SSA. Therefore, national, regional, and international agencies are now placing a high priority on developing the local rice sector in SSA as an important component of food security, national economic growth, and poverty alleviation. Rice is grown and consumed in 38 countries of SSA (Table I). Figure 3 shows the distribution of rice areas in SSA. Of the total of 8.46 million hectares (ha) of harvested rice area in SSA (5.5% of the global rice area) in

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60 A Paddy rice production (million Mg)

14 12 10 8 6 4 2 0 1961

1966

1971

1976

1981

1986

1991

1996

2001

Year Côte d’Ivoire Nigeria Sub-Saharan Africa

Kenya Senegal

Madagascar South Africa

B 8

Rice import (million Mg)

7 6 5 4 3 2 1 0 1961

1966

1971

1976

1981

1986

1991

1996

2001

Year Côte d’Ivoire Nigeria Sub-Saharan Africa

Kenya Senegal

Madagascar South Africa

Figure 2 Local paddy rice (unmilled rice) production, 1961–2005 (A) and imports of milled rice, 1961–2004 (B) into selected countries and SSA (source: FAO‐STAT, FAO, 2006a).

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Table I Harvested Rice Area, Percent of Irrigated Area, Production, and Yield of Unmilled Rice (Paddy) in African Countries (2004) Area (103 ha)a

Fully/partially irrigated (%)b

Angola Benin Burkina Faso Burundi Cameroon Central African Republic Chad Comoros Congo, DR Congo Republic Coˆte d’Ivoire Ethiopia Gambia Ghana Guinea Guinea Bissau Kenya Liberia Madagascar Malawi Mali Mauritania Mauritiusc Mozambique Niger Nigeria Re´union Rwanda Senegal Sierra Leone Sudan Swaziland Tanzania Togo Uganda Zambia Zimbabwe

20.0 33.0 51.0 19.5 20.0 14.5 80.0 14.0 415.0 2.0 500.0 8.4 16.0 119.4 525.0 65.0 11.0 120.0 1222.7 30.0 451.0 17.0 0.0 179.0 27.8 3704.0 0.04 13.0 95.0 210.0 4.8 0.05 330.0 35.0 93.0 10.0 0.25

0 2 18 21 95 – 9 – 0 0 7 85 7 4 10 1 100 2 52 28 22 100 0.0 2 80 15 100 8 50 – 75 100 27 1 2 0 –

Total/mean for SSA

8459.4

19.8

Country

a

Production (103 Mg)a 16.0 70.0 95.2 64.5 62.0 29.7 109.0 17.0 315.1 1.5 1150.0 15.5 22.0 241.8 900.0 127.0 50.0 110.0 3030.0 49.7 877.0 77.0 0.0 201.0 76.5 3542.0 0.08 46.2 264.5 265.0 15.8 0.17 647.0 68.1 140.0 12.0 0.6 12,714.0

Yield (Mg ha1)a 0.80 2.12 1.87 3.31 3.10 2.05 1.36 1.21 0.76 0.75 2.30 1.86 1.38 2.03 1.71 1.95 4.55 0.92 2.48 1.66 1.94 4.53 0.0 1.12 2.75 0.96 2.00 3.55 2.78 1.26 3.28 3.40 1.96 1.95 1.51 1.20 2.40 1.50

Source of basic data: FAO‐STAT (FAO, 2006a). Best estimates from data (1995–2004) obtained from FAO‐Aquastat (2005) at www.fao.org/ WAICENT/FAOINFO/AGRICULT/AGL/aglw/aquastat/countries/index.stm (accessed May 10, 2006) and FAO‐CORIFA (2005). c Mauritius has a few hectares of rice now; however, it plans to replace some sugarcane fields with irrigated rice in coming years to reduce rice imports. b

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30⬚

15⬚

0⬚

−15⬚ Each dot represents 2500 ha −30⬚

0

1000 (km)

−15⬚

0⬚ Figure 3

15⬚

30⬚

45⬚

60⬚

Distribution of rice areas in SSA.

2004, Nigeria and Madagascar accounted for 60% of the rice land; nine countries cultivated rice on more than 100,000 ha each; four countries grew rice on 50,000–100,000 ha each; and others were small rice producers with less than 50,000 ha each (Table I). All these countries together produced only 3% of the global rough rice output of more than 600 million Mg per annum. In 2004, rice yields varied from 0.76 Mg ha1 in Congo DR to 4.53–4.55 Mg ha1 in Mauritania and Kenya, with an average of 1.50 Mg ha1 for the whole of SSA (FAO, 2006a). In 1960–2000, rice area increased in SSA, but rice yields stagnated at a low level or decreased (Fig. 4). From 2000 to 2005, average rice yields continued to decline in West Africa and Nigeria, and increased only slightly in East Africa and Madagascar. Therefore, any increase in national rice production came from an expansion in area rather than a substantial increase in productivity. This is a disturbing trend given the escalating demand for rice in many SSA countries.

INCREASING RICE PRODUCTION IN SUB‐SAHARAN AFRICA

63

10 Sub-Saharan Africa Western Africa

Rice area (million ha)

8

Eastern Africa Madagascar Nigeria

6

4

2

0 1960

1970

1980 Year

1990

2000

Rice yield (Mg ha−1)

2.5

2.0

1.5

Sub-Saharan Africa Western Africa

1.0

Eastern Africa Madagascar Nigeria

0.5 1960

1970

1980 Year

1990

2000

Figure 4 Trends in rice area and yield in Nigeria, Madagascar, West Africa, East Africa, and SSA, 1960–2005 (source: FAO‐STAT, FAO, 2006a).

III.

WETLANDS: THE POTENTIAL RESOURCE FOR RICE PRODUCTION IN SSA

The rice plant is adapted to growing in swampy conditions and is sensitive to water stress. This makes wetlands the most suited areas for rice cultivation. The types and distribution of wetlands and their potential for rice cultivation in SSA are discussed in this section.

64

V. BALASUBRAMANIAN ET AL.

A. DEFINITION, AREA, AND DISTRIBUTION

OF

WETLANDS

For the purpose of agricultural land‐use planning, wetlands can be defined as areas subject to periods of completely water‐saturated soils with possibilities of flooding during part of the crop‐growing period. The depth and duration of flooding of these soils depend on the position of the land on the catena and the extent of drainage available; the flooding may be for part of the growing season or throughout the year. Wetlands are found not only in low‐lying areas (river and coastal floodplains, deltas, and depressions) but also on upper river terraces, foot slopes, and hilltops. For example, in West and Central Africa, wetlands are a part of the inland valley system, which is a continuum of drylands on upper and mid‐slopes and wetlands in valley bottoms. On lower slopes, the boundary between wetland and dryland is often gradual. The distribution of normal and salt‐aVected (saline) wetlands along with major rivers and lakes in SSA is shown in Fig. 5. Our discussion on area, types and characteristics, and distribution of wetlands in SSA is derived from Andriesse (1986). On the basis of the 1:5 million soil map of FAO (1977), the estimated wetland area in SSA is 239 million ha (Table II) (Andriesse, 1986). About 40% of the total wetland area is found in the equatorial region and the rest in the Guinea savannas (Fig. 5). About 36 million ha of wetlands are located in the tropical highlands of East and Central Africa and Madagascar. These estimates of wetland area must be used cautiously because of the possible large margins of error in estimates derived from small‐scale maps.

B. TYPES

AND

CHARACTERISTICS

OF

WETLANDS

Wetlands in SSA are grouped into four categories: inland basins and drainage depressions, inland valleys, river floodplains, and coastal wetlands. A brief description of each category is given below. 1.

Inland Basins

Inland basins are the largest area of wetlands potentially suitable for rice cultivation, they occupy 108 million ha (45% of the total wetland area in SSA) (Table II). They comprise drainage depressions and inland deltas of rivers, with imperfectly to poorly drained and potentially acidic soils (Ultisols, Oxisols, Alfisols, Entisols, and Vertisols). Examples are the Upper Nile, Sokoto, Lake Chad, and Congo basins; the shallow swamps around Lake Bangweulu and the Kafue flats; and the lacustrine deposits of Lake Rukwa and Lake Eyasi.

INCREASING RICE PRODUCTION IN SUB‐SAHARAN AFRICA

65

30⬚

15⬚

0⬚

−15⬚

Wetlands −30⬚

Saline wetlands Rivers and lakes

0

1000 (km)

−15⬚ Figure 5 SSA.

0⬚

15⬚

30⬚

45⬚

60⬚

Distribution of normal and saline wetlands along with major rivers and lakes in

Table II Estimated Areas of Four Types of Wetlands in SSAa

Area (million ha)

Percentage of total wetlands

Percentage of total area

Inland basins/depressions Inland valleys River floodplains Coastal wetlands

107.5 85.0 30.0 16.5

45 36 12 7

9.0 7.0 2.5 1.5

Total

239.0

100

20.0

Type of wetlands

a

Adapted from Andriesse (1986).

V. BALASUBRAMANIAN ET AL.

66

2.

Inland Valleys

Inland valley wetlands occupy 85 million ha (36% of the total wetland area in SSA) (Table II); only 10–15% of the inland valley area is used for agriculture. Narrow inland valleys are located upstream from river floodplains that are much wider. Each inland valley represents a toposequence of a valley bottom with its hydromorphic edges, and the contiguous dryland slopes and crests that contribute runoV and seepage to the valley bottom. They are known as dambos or boliland in eastern and central Africa, fadamas in northern Nigeria and Chad, bas‐fonds or marigots in francophone Africa, and inland valley swamps in Sierra Leone. Most inland valley wetlands are concentrated in the intertropical zone where rainfall is superior to 700 mm, and their catchment sizes generally range from 100 to 2000 ha. The soils (Entisols) in the valley bottoms are flooded during the rainy season, whereas the soils (Ultisols, Oxisols, Alfisols, and Inceptisols) on adjacent drylands are aerobic and erosion‐prone.

3.

River Floodplains

An estimated area of 30 million ha (12% of the total wetland area) is under river floodplains in SSA (Table II). A floodplain is a wide, flat plain of alluvium bordering streams and rivers that flood it periodically. Well‐developed floodplains extend from tens of meters to tens of kilometers on either side of large rivers such as the Gambia, Niger, Benue, Zaire, Zambezi, Limpopo, Tana, White and Blue Nile, and Chari. Soils of the floodplains (Entisols, Inceptisols) are moderately well to poorly drained and medium‐ to fine‐ textured with moderate to high fertility. Soils can be saline and/or alkaline in drier regions.

4. Coastal Wetlands Coastal wetlands cover an estimated area of 16.5 million ha (7% of the total wetland area in SSA) (Table II). They comprise Deltas (e.g., Niger in Nigeria, Rufiji in Tanzania, and Zambezi in

Mozambique) Estuaries (at the mouths of the Zaire, Cross, Gambia, and Corubal rivers) Intertidal flats or lagoons along the West and East African coasts.

Soils (Entisols, Inceptisols, Histosols) are poorly drained and nonsaline in freshwater swamps, acid sulfate (Entisols, Inceptisols) in mangrove swamps,

INCREASING RICE PRODUCTION IN SUB‐SAHARAN AFRICA

67

poorly drained and saline (Inceptisols) in lagoons, coarse‐textured (Entisols, Inceptisols) in sand bars and dunes, and organic (Histosols) in permanently flooded areas.

IV. RICE SOIL RESOURCES Soils provide the base for crop production. The physical and chemical characteristics of soils reflect the parent materials from which they are derived. For example, coarse infertile soils are developed from poor and acid rocks (sandstones, granites, quartzites, rhyolites), coarse‐ to medium‐ textured soils of low to moderate fertility from intermediate rocks (hornblende, granites, quartz‐feldspar gneisses, diorites, andesites), and medium to fine, relatively fertile soils from rich parent rocks (amphibolites, dolerites, basalts, hornblende gneisses, shales, siltstones). Soils derived from volcanic ash are rich in most nutrients except N and P. In addition, the climate, the extent of weathering, the position of soils on landscape, vegetation cover, and soil moisture regimes aVect the physical, chemical, and biological properties of soils. In this section, we briefly discuss the two broad categories of soils—dryland and wetland—and their characteristics.

A. DRYLAND SOILS AND THEIR CHARACTERISTICS Dryland soils on the upper and middle slopes of the catena are generally well drained, deep to very deep, and coarse‐ to medium‐textured or gravelly in the humid forest and Guinea savanna zones. Soils in the upper parts of the toposequence are moderately to well drained, shallow to medium in depth, and coarse‐ to medium‐textured or gravelly in the Sudan savanna transition zones (Andriesse and Fresco, 1991). Most of the dryland soils are highly weathered and greatly leached with low‐activity clays (silicious, kaolinitic, and halloysitic minerals) in the humid zones and less leached and fairly rich in weatherable minerals in the Sudan savanna transition zones. The red clayey or clay‐loam soils dominant in the humid forest and Guinea savanna zones are classified as Ultisols and Oxisols, and some as Alfisols; the brown sandy or sandy loam soils of the Sudan savanna transition zones belong to Alfisols, Inceptisols, or Entisols; and the organic matter (OM)‐ rich, relatively fertile soils of the highlands of East and Central Africa and Madagascar are mostly Inceptisols. Volcanic ash soils found in some dryland rice areas (e.g., in parts of Cameroon) are also Inceptisols (Kawaguchi and Kyuma, 1977; Sanchez and Buol, 1985). Gravelly soils (Entisols) and soils on steep slopes are highly prone to erosion and not suitable for dryland rice.

V. BALASUBRAMANIAN ET AL.

68

The inherent fertility of most dryland soils is low (low pH, cation exchange capacity, and base saturation) in warm and humid zones due to intense weathering of parent materials, and medium to high in semiarid zones due to less intense weathering and, in West Africa, due to deposition of Ca, Mg, and K from dust deposits from the Sahara Desert that occur in the dry season (DS) (Andriesse and Fresco, 1991). Subsoil acidity and Al and Mn toxicity are common problems in some soils. Soil OM and N contents decrease as we move from humid to subhumid to semiarid zones. Total soil P is generally low, with only 2–4% of the total P available to plants (Table III) (Breman, 1998). This is because of the high P‐fixing capacity of fine‐ textured soils found in humid and subhumid zones (Abekoe and Sahrawat, 2001; Juo and Fox, 1977). Therefore, P fertilization is an important requisite to get a crop response to other nutrients such as N. K deficiency is more severe in humid regions than in drier areas.

B. WETLAND SOILS

AND

THEIR CHARACTERISTICS

Rice soils converted from natural wetlands are fairly rich in exchangeable bases (Ca, Mg, and K), slightly acidic to neutral (pH 6–7), low in P‐fixation capacity, and not Al toxic. The soils are often mineral in seasonally dry wetlands and peaty in permanently flooded wetlands. Most mineral wetland soils are fluvial, lacustrine, estuarine, or marine alluvial materials. In soil taxonomy (Soil Survey StaV, 1998), soils with an aquic moisture regime and specified morphologic characteristics of wetness are distinguished at the suborder level (e.g., Inceptisols: Aquepts; Alfisols: Aqualfs). At the subgroup level, aquic subgroups with signs of wetness only in lower horizons are not considered as wetland soils, except the aquic subgroups of flooded Vertisols,

Table III Mean Chemical Characteristics of Dryland Soils (0–0.3 m) in Four Agroecological Zones of West Africa Soil properties pH (water) Org. C (g kg1) Total N (g kg1) Total P (mg kg1) Bray‐1 P (mg kg1) CECa (cmol kg1) Base saturation (%)

Equatorial forest

Guinea savanna

Sudan savanna

Sahel

5.7 20 2.0 260 9 8.7 28

5.7 12 1.3 340 7 8.5 59

6.7 6 0.5 210 4 8.1 69

5.7 3 0.2 100 4 2.5 28

a CEC, cation exchange capacity. Adapted from Breman (1998).

INCREASING RICE PRODUCTION IN SUB‐SAHARAN AFRICA

69

such as Chromoxererts, Chromuderts, Pelloxererts, and Pelluderts, with distinct mottles in the upper 0.5 m of the soil profile. Histosols are mostly wetland soils. In another classification of soils, the FAO–UNESCO legend for the soil map of the world (FAO, 1977), the Gleysol, Fluvisol, Planosol, and Histosol and gleyic units of Acrisol, Arenosol, Cambisol, Luvisol, Podsol, and Regosol and periodically wet Ferralsol refer to wetland soils. Major wetland rice soils of SSA, their classification in soil taxonomy and the FAO–UNESCO system, and a third system based on fertility capability classification (FCC) and their potential for rice cultivation are given in Table IV. The FCC (Buol et al., 1975) is a technical system for grouping soils according to the kinds of problems they present in agronomic management of crops. It is based on quantitative topsoil and subsoil (up to 0.5 m) parameters directly relevant to plant growth. Additional locally important features can be considered and added to the FCC as primes (0 ) or asterisks (). The adapted FCC of wetland soils (Sanchez and Buol, 1985) is used for the FCC in Table IV. Rice soils from converted natural wetlands include alluvial (Entisols), hydromorphic (Inceptisols), black cracking clay (Vertisols), and red clayey (Ultisols and Oxisols) or loamy soils (Oxisols or Alfisols). Except for the black clay soils, they are acidic, moderately to highly P‐fixing (Abekoe and Sahrawat, 2001), and low in potentiality, with less than 10% weatherable minerals; they are also diYcult to puddle. These soils scattered all over SSA are moderately to highly suitable for rice cultivation (Table IV). Organic soils (OM > 20%) or Histosols include peat and muck soils found in continuously flooded wetlands in low‐ and high‐altitude regions (forest zones of West Africa and highlands of East and Central Africa and Madagascar). There are two classes of organic soils: deep (organic in top 0.5 m of the profile) and shallow organic over (clay, loamy, or sandy) mineral soils within the top 0.2 m of the profile. Shallow organic soils over clay are the most common in SSA and elsewhere, and they are low in bearing capacity (mechanization not possible) and deficient in many nutrients (Sanchez and Buol, 1985). Acid sulfate soils are organic soils over clay with marine sediments found in some coastal wetlands of SSA. Many soils with adequate water retention capacity but lacking an aquic soil moisture regime are converted into artificial (or anthropic) wetlands through terracing and irrigation. Two important soil types of artificial wetlands are oxic soils (Oxisols, Oxic Ultisols/Alfisols) and Andepts (volcanic ash soils). Oxic soils have more than 35% clay, low pH and cation exchange capacity, high levels of free iron that can result in Fe toxicity when flooded, and high P‐fixing capacity and Al toxicity (Abekoe and Sahrawat, 2001; Sahrawat, 2004a); they are generally diYcult to puddle and have low to moderate potential for rice cultivation (Sanchez and Buol, 1985). These soils are found in rainforest and derived savanna zones. Volcanic ash soils

V. BALASUBRAMANIAN ET AL.

70

Table IV Wetland Rice Soils, Their Classification, and Their Potential for Rice Production in Africa

Soil type

Soil Survey StaV FAO–UNESCO Soil (1998) Map (FAO, 1977)

A. Wetland rice soils converted from natural wetlands Young alluvial Entisols: Fluvisols: Dystric, soils Ustifluvents, Eutric Torrifluvents, Arenosols: Gleyic Xerofluvents, Ustipsamments Black or peaty Inceptisols: Gleysols: Humic soils Epiaquepts, Humaquepts Hydromorphic Vertisols: Vertisols: Pellic Vertisols Calciaquerts, Duraquerts, Dystruderts, Calcixererts, Durixererts Inceptisols: Epiaquepts Red clay soils: Ultisols: Udults, Acrisols: Ferric, humid zone Aquults Gleyic Red clay‐loam Oxisols: Aquox, Ferralsols: Orthic, soils: Ustox humic Luvisols: Orthic, subhumid and Alfisols: Aqualfs, Ferric semiarid zones Udalfs, Ustalfs

FCCa

SLgek, SCghek

Saline soils

Inceptisols: Epiaquepts, Aridisols: Natrargids

High

Lgak (coastal), Moderate Cgaik (highlands) Cvg

High

Cgaeik0

Moderate to low Moderate

(LC)ghk0 d, Lgh(e)k

B. Wetland rice soils in anthropic (artificial) wetlands on terraced landscapes Oxic soils Oxisols, Oxic Ferralsols, Acrisols, Cigae (>35% clay) Ultisols/Alfisols Luvisols Tropical brown Inceptisols: Cambisols: Gleyic, Lxd, Lxg soils/volcanic Epiaquepts, Eutric ash soils Sulfaquepts, Andosols: Ochric, Halaquepts Humic Entisols: Hydraquents C. Problem soils Acid sulfate soils Entisols: Sulfaquents Inceptisols: Sulfaquepts Peats (deep) Histosols (>20% organic matter) Peats (shallow)

Potential for rice cropping

Low/ moderate High/very high

Fluvisols: Thionic

OCcgak

Low/ moderate

Histosols: (>20% organic matter)

Ogakk0

Low/very low

OCgakk0 , OLgakk0 , OSgakk0 Ssgbkk0 , SCsgbkk0

Low/ moderate

Solonchaks: Gleyic

Limited

INCREASING RICE PRODUCTION IN SUB‐SAHARAN AFRICA

71

Table IV (continued)

Soil type Alkaline soils

Degraded gray soils

Soil Survey StaV FAO–UNESCO Soil (1998) Map (FAO, 1977) Inceptisols: Solonets: Gleyic Natraquepts Alfisols: Natraqualfs Vertisols: Natraquerts Ultisols: Acrisols: Plinthic Plinthaquults

Skeletal soils Ultisols: (>35% gravel) Endoaquults

Planosols: Eutric, Dystric

FCCa Cngbgkk0 , SCngbkk0

Potential for rice cropping Limited

SLgkk0 ea, CLgkea

Moderate (H2S toxicity) SL00 gkk0 i, LC00 gkk0 i, Very low SRgk, LRgk

a

FCC, fertility capability classification (Sanchez and Buol, 1985): S, sand; C, clay; L, loam; O, organic; R, rocky; c, acid sulfate; x, X‐ray amorphous allophonic; g, gleic; g, pergleic; gd, wetland soils dry for 60 days or more; k, low potential productivity with <10% weatherable minerals; k0 , low in K; e, low cation exchange and buVering capacity; a, acid pH < 5 þ Al toxic þ P‐fixing; h, acid, pH 5–6 but not Al toxic; b, basic pH > 7.3; v, vertic; s, saline; n, alkaline; %, steep slope; i, high levels of free iron, highly P‐fixing. Note: FCC ratings for diVerent soils are first approximations based on available information, and they can be modified or further improved with additional data at local level.

(Inceptisols) are dominant in X‐ray amorphous allophone minerals, rich in all nutrients except N and P, and high in P‐fixation capacity; they pose moderate diYculty to puddling and regenerate the structure easily on drying. They are good for rice dryland crop rotations with high productivity (Kawaguchi and Kyuma, 1977). DiVerent types of problem soils occur in SSA: acid sulfate soils (Inceptisols: Sulfaquepts) in mangrove swamps, deep and shallow peats (Histosols) in low‐ and high‐altitude areas, saline (Inceptisols: Halaquepts) and alkaline (Inceptisols: Natraquepts; Vertisols: Natraquerts) soils in the Sahel and coastal wetlands, degraded soils (Ultisols: Aquults) in the humid forest zones, and skeletal soils (Ultisols: Aquults) on adjacent sloping drylands of inland valleys. Soil pH decreases drastically and Al toxicity increases only when acid sulfate soils are drained. Remediation of adverse conditions and the combined use of salt‐tolerant rice varieties and integrated nutrient application are needed to grow rice on saline and alkaline soils.

V. AGROCLIMATIC ZONES AND RICE ECOSYSTEMS The suitability of land for various crops is determined by climate and weather variables (agroclimatic zones), landscape moisture regimes (physio‐ hydrographic positions), and soil characteristics. Rice thrives in areas with

V. BALASUBRAMANIAN ET AL.

72

warm temperatures (above 20 C during the cropping season), annual rainfall ranging from 0.5 to >1.5 m, and growing periods of 90þ days. Rainfall distribution is generally monomodal with a distinct humid period in areas with a rainfall range of 500–1000 mm per annum (e.g., northern and southern Guinea savanna zones) and bimodal with two growing seasons toward the equator (Andriesse, 1986). The classification of land and water resources into (agro)ecosystems is helpful for the planning, development, management, and monitoring of these resources. The choice of classification scheme is always subjective to some extent and will depend on the purpose of classification. Andriesse and Fresco (1991) defined 18 rice environments by combining six landscape or physio‐hydrographic positions with three agroecological zones (AEZs) for West and Central Africa. For this chapter, we use surface‐water regimes to define four major rice ecosystems: dryland, rainfed wetland, deepwater and mangrove swamps, and irrigated wetland (Fig. 6). The distribution of rice areas into four ecosystems and the related environmental/human disease constraints for diVerent African countries are shown in Table V. For 16 large rice‐growing countries, actual mean rice areas (1995–2004) in diVerent ecosystems are provided in Table VI.

A. DRYLAND RICE ECOSYSTEMS Dryland rice is also known as ‘‘upland’’ or ‘‘pluvial’’ rice. It is cultivated on level or sloping lands and on hydromorphic fringes (Fig. 6) in fields that do not have bunds to retain surface water (Sie, 1991). Flooding is rare in this ecosystem, and dryland rice depends solely on rainfall and should have a water table that remains at 0.5 m or more below the soil surface. Similar to Plateau Slopes Hydromorphic edge Valley bottom

Floodplains

Dryland DW/Mangrove Rainfed wetland Irrigated wetland Figure 6 Major rice ecosystems of SSA as defined by surface‐water regimes.

INCREASING RICE PRODUCTION IN SUB‐SAHARAN AFRICA

73

Table V Rice Area by Ecology/Ecosystem and Related Environmental and Human Disease Problems for Countries of SSA

Country

Percentage of area by Mean area ecology/ecosystema (1995–2004) (103 ha)a IRb RWRc DRd DWRþe

Angola Benin Burkina Faso

21.2 21.7 47.9

0 2 46

100 7 50

0 91 4

0 – 0

Burundi Cameroon Central Africa Republic Chad Comoros Congo, DR

17.7 16.9 13.1

21 95 0

74 5 3

4 0 97

0 0 0

83.7 14.0 445.2

9 – 0

6 – 11

0 – 89

85 – 0

1.8 500.9 8.4

0 7 85

100 15 15

0 78 0

0 0 0

0.5 13.9 116.9 495.5

100 7 24 10

0 63 21 23

0 16 55 44

0 14 0 23

Guinea Bissau

65.5

1

27

27

45

Kenya

11.8

100

0

0

0

Liberia Madagascar Malawi Mali Mauritania Mauritius Mozambique

121.0 1195.7 43.8 365.3 18.1 0 164.6

2 52 28 22 100 0 2

6 18 72 13 0 0 59

92 29 0 1 0 0 39

0 1 0 64 0 0 0

Niger Nigeria Re´union Rwanda

22.7 2466.4 0.04 5.1

80 17 100 8

0 47 0 92

0 30 0 0

20 6 0 0

79.2 236.6 4.7

50 0 75

40 28 –

0 68 –

10 4 –

Congo Republic Coˆte d’Ivoire Ethiopia (2001–2004) Gabon Gambia Ghana Guinea

Senegal Sierra Leone Sudan

Environmental and human disease problems HIV, malaria, cholera Malaria Bacteria/N/Cl1 pollution, salinity Malaria, bilharzia Malaria, bilharzia Malaria Poor water and sanitation Malaria/diarrhea/dengue Petroleum, mineral pollution – Malaria/worms Salinity, malaria – Salinity, deforestation Water pollution, malaria Malaria/diarrhea/bacterial diseases Salinity/silting, malaria/diarrhea Land/wildlife park conflicts, malaria – Malaria, bilharzia Malaria Bilharzia Salinity, soil degrading – Salinity, HIV/malaria/ diarrhea/bilharzia Malaria/diarrhea/bilharzia Pollution, malaria/worms – Deforestation/soil erosion/ peat drying Acid/alkalinity/pollution Deforestation/peat harvest Silting/pollution (continued)

74

V. BALASUBRAMANIAN ET AL. Table V (continued)

Country Swaziland Tanzania Togo Uganda Zambia Zimbabwe

Total/mean for SSA

Percentage of area by Mean area ecology/ecosystema (1995–2004) (103 ha)a IRb RWRc DRd DWRþe 0.05 439.9 38.1 71.0 10.9 0.2

7180.0

100 4 1 2 0 –

19.8

0 73 19 53 100 –

0 23 80 45 0 –

33.5

37.8

0 0 0 0 0 –

8.9

Environmental and human disease problems Malaria, bilharzia Salinity/pollution/silting Pollution, malaria Salinity/silting, malaria Malaria/diarrhea Civil strife, malaria/ bilharzia/diarrhea/ agrochemical poisoning –

a

Best estimates from data (1995–2004) obtained from FAO‐Aquastat (2005) at www.fao.org/ WAICENT/FAOINFO/AGRICULT/AGL/aglw/aquastat/countries/index.stm (accessed May 10, 2006), FAO‐CORIFA (2005), Defoer et al. (2002), Adegbola and Singbo (2005), and Ezedinma (2005). b IR, irrigated wetland rice. c RWR, rainfed wetland rice. d DR, dryland rice. e DWRþ, deepwater and mangrove rice.

the situation in Asia, dryland rice is generally a subsistence crop in Africa. It is of critical importance for the local food security of poor communities that do not have access to wetland fields. In Coˆte d’Ivoire, Guinea, Guinea Bissau, and Sierra Leone, dryland rice is the only staple available between the maize and cassava harvests or between sweet potato and cassava (McLean et al., 2002). In the Zambezi region of Mozambique, dryland rice is the main source of food for the poor.

1. Dryland Rice Area Of the global dryland rice area of 14 million ha in the 1990s, 2.7 million ha are planted to dryland rice in Africa (Table III); it accounts for an estimated 38% of the total rice area in SSA. Nearly 97% of the cultivated rice area in the Central African Republic is in drylands. Other countries with dominant dryland rice ecosystems (more than 65% of the cultivated rice area) are Liberia, Benin, Congo DR, Togo, Coˆte d’Ivoire, and Sierra Leone (Table V). Countries with more than 100,000 ha of dryland rice are Congo DR, Guinea,

INCREASING RICE PRODUCTION IN SUB‐SAHARAN AFRICA

75

Table VI SSA Countries with Large Areas in Rice Ecologies/Ecosystems (Mean of 1995–2004) 1995–2004 Mean rice area (103 ha)a Country Benin Burkina Faso Chad Congo, DR Coˆte d’Ivoire Ghana Guinea Liberia Madagascar Mali Mozambique Nigeria Senegal Sierra Leone Tanzania Togo Uganda

IRb

RWRc

DRd

DWRþe

Total

0.4 22.0 7.5 0.0 35.1 28.1 49.5 2.4 621.8 80.4 3.3 419.3 39.6 0.0 17.6 0.4 1.4

1.5 24.0 5.0 49.0 75.1 24.5 114.0 7.3 215.2 47.5 97.1 1159.2 31.7 66.2 321.1 7.2 37.6

19.8 1.9 0.0 396.2 390.7 64.3 218.0 111.3 346.8 3.6 64.2 739.9 0.0 160.9 101.2 30.5 32.0

0.0 0.0 71.2 0.0 0.0 0.0 114.0 0.0 12.0 233.8 0.0 148.0 7.9 9.5 0.0 0.0 0.0

21.7 47.9 83.7 445.2 500.9 116.9 495.5 121.0 1195.7 365.3 164.6 2466.4 79.2 236.6 439.9 38.1 71.0

a

Best estimates from data (1995–2004) obtained from FAO‐Aquastat (2005) at www.fao.org/ WAICENT/FAOINFO/AGRICULT/AGL/aglw/aquastat/countries/index.stm (accessed May 10, 2006), FAO‐CORIFA (2005), and FAO‐STAT (FAO, 2006a). b IR, irrigated wetland rice. c RWR, rainfed wetland rice. d DR, dryland rice. e DWRþ, deepwater and mangrove rice.

Coˆte d’Ivoire, Liberia, Madagascar, Nigeria, Sierra Leone, and Tanzania (Table VI).

2. Cropping Systems Dryland rice systems range from shifting to permanent cultivation. Shifting or slash‐and‐burn cultivation is common in humid forest zones of West Africa where farmers cut and burn the bush‐fallow vegetation and plant rice as the first crop to exploit the soil fertility built up during the fallow period and the ash from burning. They may apply a little manure or compost, but no chemical fertilizers are used. As this practice depletes the soil, weeds build up and rice yields decline drastically after the second crop, when farmers plant cassava on old plots and move to a new area for rice cultivation.

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Nutrient mining degrades the soil in this type of slash‐and‐burn system (Fernandez et al., 2000; Oldeman et al., 1991). Farmers plant rice as a sole crop or mixed with maize, beans, yam, cassava, or plantains (mixed cropping) to avoid risks. In areas with a long growing season and suYcient rainfall, dryland rice is rotated with maize, cowpea, beans, soybean, or sweet potato. In inland valleys, dryland rice and cash crops are often grown on hydromorphic fringes while fodder crops and trees grow on upper slopes and crests. In the highlands of East Africa, dryland rice is rotated with wheat, maize, potato, or sweet potato.

3.

Cultivation Practices and Yields

In this ecosystem, the fields are not bunded, there is no flooding, and the soil remains aerobic (not saturated with water) for most of the growing season. Rice seeds are sown by broadcasting or dibbling in hand‐hoed fields. An adequate supply of soil water is critical for good plant growth and yield. This can be achieved by in situ rainwater harvesting (RWH) through improved infiltration of rainwater by proper tillage, reduced water loss from the soil surface by proper mulching or plant cover, and improved crop‐water use by selecting adapted varieties and following moisture‐conserving cultivation practices (Hatibu et al., 2000). As dryland rice farmers are generally poor, they may apply some household wastes and other organic manure but do not generally apply purchased inputs to their rice crops. At harvest time, mature panicles are collected, dried, and threshed manually. Dryland rice yields range from less than 0.5 Mg ha1 on subsistence farms to 2 Mg ha1 in well‐managed permanent cropping of rice in rotation with legumes and other crops. Yields are low on subsistence farms because of poor cultivation methods, low‐input use, excessive weeds, and depleted soils. With the introduction of dryland NERICA varieties in SSA, farmers in Uganda obtained average rice yields of 2.2 Mg ha1 with moderate inputs (Kijima et al., 2006).

4.

Production Constraints

Both abiotic and biotic constraints limit rice production in drylands. Serious abiotic constraints include variable rainfall, low temperature in high‐altitude areas, and poor soils. The total rainfall of 0.9–2.0 m is adequate in the humid and subhumid areas where dryland rice is generally grown in SSA, but the rainfall distribution can be poor, with unpredictable dry spells. On the other hand, the temperature during the growing season is

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relatively favorable in most parts, except in the tropical highlands of East and Central Africa and Madagascar. Degradation of soil structure and surface sealing constrain crop emergence and growth in semiarid areas (Andriesse and Fresco, 1991). Dryland rice requires extensive weeding to get a decent yield and, in such systems, topsoil erosion can be a serious problem, especially on slopes. Drought is another serious problem for dryland rice due to the inadequate quantity and/or poor distribution of rainfall and shallow depth of and surface crusting in some soils. Among biotic factors, weeds are the most serious, followed by blast and brown spot diseases. Weeds are generally more competitive than rice in infertile dryland soils. Striga is a parasitic weed of dryland cereals, including rice. Estimated yield losses due to weeds range from 30 to 100%. Weed infestation and loss of N reduce yields by 25% on intensive dryland rice farms of West Africa (Becker and Johnson, 2001). Stem borers and rice bugs are the major insect pests. Nematodes are serious problems in continuously monocropped dryland rice fields (Coyne et al., 2004; Plowright and Hunt, 1994) and can reduce yields by up to 30%. Termites are problems in some areas. Rodents and birds damage rice crops in all ecosystems.

B. WETLAND RICE ECOSYSTEMS Unlike the dryland ecosystem, rice fields in the wetland ecosystem are flooded during the growing season. We distinguish three types of wetland rice ecosystems—rainfed wetland, deepwater, and irrigated—as determined by the surface‐water regime. The ecosystem is considered to be rainfed wetland when the water supply to crops is from rainfall and groundwater. In contrast, in the deepwater ecosystem, most of the water on the fields is from the lateral flow of water onto the land. In the irrigated wetland ecosystem, a significant part of the water supply is from irrigation.

1. Rainfed Wetland Rice Ecosystem Rainfed wetland rice is grown on lower parts of the toposequence and in valley bottoms (Fig. 6) in level to slightly sloping bunded fields that are flooded by rainwater for a part of the growing season to water depths that may exceed 1.0 m for not more than 10 consecutive days. Both rainwater and stored groundwater support rainfed wetland rice. Four types of flooded wetlands are recognized in Africa: riverine shallow, riverine deep, boliland (grassy inland swamps), and mangrove (Andriesse and Fresco, 1991; Buddenhagen, 1986).

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Rainfed wetlands are characterized by a lack of water control, with droughts and floods being potential problems (Hatibu et al., 2000; McLean et al., 2002). On the basis of the constraints, rainfed wetlands can be divided into four subecosystems: (1) favorable, (2) drought‐prone, (3) submergence‐ prone, and (4) drought‐ and submergence‐prone. All four subecosystems occur in SSA. Most inland valley bottoms represent the favorable rainfed wetlands which could be the next frontier for intensification of rice and rice‐ based cropping systems. Rice varieties and production technologies developed for the irrigated ecology can be easily adapted for rice in favorable rainfed wetlands. Other rainfed wetland (drought‐, flood‐, and drought‐ and flood‐prone) subecosystems are found in riverine shallow, riverine deep, hydromorphic edges of inland valleys, and mangrove ecologies. They are highly diverse, with often variable rainfall patterns, adverse soils, and many abiotic and biotic constraints. In addition, farmers are poor and have to adapt their cropping practices to the complex risks, potentials, and problems characteristic of such ecosystems. a. Rainfed Wetland Rice Area. Of the global rainfed wetland rice area of 54 million ha, an estimated 2.4 million ha are found in SSA; this is equal to an estimated 33% of the total rice area in SSA (Table V). The rainfed wetland rice ecosystem is generally found in the humid and subhumid forest and moist savanna zones of SSA. During 1995–2004, Nigeria had the largest area under rainfed wetland rice (1,159,208 ha), followed by Tanzania (321,127 ha), Madagascar (215,226 ha), Guinea (113,965 ha), Mozambique (97,114 ha), Cote d’Ivoire (75,135 ha), and Sierra Leone (66,248 ha) (Table VI). b. Cropping Systems. The number and choice of rice and other crops to grow on rainfed wetlands depend on the length of the rainy season and water availability. Farmers generally grow one crop of rice during the wet season (WS) and leave the land fallow for the rest of the year. Vegetables, sweet potato, or taro may be grown on mounds or bunds adjacent to flooded rice fields (Andriesse and Fresco, 1991). If the rains extend to 5 months or more, farmers can grow a post‐rice crop such as corn, soybean, vegetables, or wheat (in highlands). Beans, cowpea, and vegetables are grown with residual moisture in many inland valley bottoms and swamps of West Africa (WARDA, 2002). c. Cultivation Practices and Yields. Farmers sow seeds by broadcasting on plowed land and fields are bunded to collect rainwater. They weed, redistribute seedlings to ensure uniform crop stand, and harvest by hand, but they add few or no purchased inputs (fertilizers or biocides). Harvested rice is threshed on dry ground in the field or near the house. Rice is dehusked by hand pounding in a mortar or by small village mills. Rice is stored in

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large baskets for home consumption or in community warehouses for sale. In West Africa, rice husks are mixed with clay and used for plastering house walls or for making mud bricks (Murray, 2005). Small farmers often plant traditional rice varieties and apply very few external inputs. As a result, they obtain yields from 1 to 2 Mg ha1 in SSA in contrast to the world average yield of 2.3 Mg ha1 for rainfed wetland rice. However, WARDA (West Africa Rice Development Association, The Africa Rice Center) has identified 46 promising rice varieties with target yields of 3 Mg ha1 or more in rainfed lowlands when moderate inputs are applied (WARDA, 2002). These high yields have been recorded in some areas in West Africa (Sakurai, 2006). d. Production Constraints. Poor water control is a major constraint to rice intensification in this ecosystem. Many abiotic and biotic constraints aVect wetland rice production. Abiotic stresses include variable rainfall, with drought and flood occurrences in the same season. Rainfed wetland rice is adversely aVected by Fe, Al, and Mn toxicity in wet forest zones (Buddenhagen, 1986; Sahrawat, 2004a) and in poorly drained soils of coastal wetlands. Inland salinity and alkalinity are problems in drier and desert areas (Van Asten et al., 2003). Weeds are the principal biotic constraint, followed by insect pests (stem borers, African rice gall midge, AfRGM, and rice bugs) and diseases (blast and brown spot) (WARDA, 1998, 2000). Rice yellow mottle virus (RYMV) is a major scourge of wetland rice and can sometimes lead to total crop failure (WARDA, 2000). In addition, rats and birds are serious problems in all ecosystems.

2.

Deepwater and Mangrove Rice Ecosystems

The deepwater ecosystem (Fig. 6) covers several environments where rice is planted, which is adapted to increasing water depths of 1.0 m or more for durations of 10 days to 5 months. These rice plants must have the ability to elongate rapidly to stay above the water surface. ‘‘Floating rice’’ can elongate up to 5 m and form adventitious roots that can absorb nutrients directly from the floodwater in addition to regular roots grounded in the soil. No varieties are available that are adapted to rapid or irregular rise of floodwater or sediment‐laden floodwater that can cover crops for longer than 10 days in some deepwater areas (McLean et al., 2002). In low‐lying coastal areas, we can diVerentiate perennially fresh, seasonally saline, and perennially saline tidal wetlands where rice plants are subject to daily tidal submergence. Plants in tidal lands do not elongate greatly, but tillering and tiller survival may be reduced in saline soils. In problem soils (acid sulfate and sodic or alkaline soils), excess water accumulates in fields due

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to poor drainage, but no prolonged submergence occurs (McLean et al., 2002). Al or Fe toxicity is a serious risk when acid sulfate soils of coastal wetlands are drained (Sahrawat, 2004a), whereas deep peat soils constrain rice production in high‐altitude areas. a. Deepwater and Mangrove Swamp Rice Area. Worldwide, about 11–14 million ha will come under deepwater ecosystems. In SSA, about 0.63 million ha are estimated to be aVected by excess flooding, tidal submergence, saltwater intrusion, and salinity and acid sulfate soils; these ecosystems cover an estimated 9% of the total rice area in SSA (Table V). Some parts of the floodplains of the Niger River, the low‐lying wetlands of Madagascar, and the poorly drained inland basins of Chad, Guinea, Mali, Niger, and Nigeria have deep flooding, whereas the low‐lying coastal wetlands of East and West Africa are aVected by salinity and alkalinity due to seawater intrusion. Mangrove swamps constitute about 49% of the rice land in Guinea Bissau, 14% in the Gambia, and 13% in Guinea (Defoer et al., 2002). b. Cropping Systems. Only rice is grown in the rainy season. The type, depth, and duration of flooding determine the rice varieties grown. Specific rice varieties with diVerent elongation ability are selected for deepwater (about 1.0 m) and very deepwater (1–5 m) conditions. When flooded, deepwater rice varieties can elongate 0.02–0.03 m per day, while floating rice varieties elongate rapidly up to 0.2 m per day (McLean et al., 2002). Non‐elongation‐type rice varieties with submergence tolerance (Xu et al., 2006) are needed for freshwater tidal wetlands where flash floods are of short duration (less than 2 weeks), whereas salt‐tolerant varieties should be selected for coastal saline lands, including mangrove swamps. In the drier mangrove areas (e.g., Casamance in Senegal, Guinea Bissau, and the Gambia) of West Africa, farmers plant rice on ridges to mitigate the problems of Fe toxicity and salinity (WARDA, 2002). During the dry or winter season, vegetables or legumes can be grown in deepwater areas. In salt‐aVected areas, the lands are too dry or saline for cropping in the DS. c. Cultivation Practices and Yields. Deepwater rice and floating rice are mainly grown in deepwater areas. With the onset of rains, the land is plowed and harrowed and dry rice seeds are sown by broadcasting. With the moisture from rainfall, the seeds germinate and the seedlings start growing. If early rainfall is regular and adequate for land preparation by puddling, seedlings raised in nurseries are transplanted. Generally, flooding occurs in the later stages of plant growth and can last for several months. Crop survival and productivity depend on the age of the rice crop when inundation starts, the rate of rise of

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the water, and the depth and duration of the flood. Sudden or flash flooding can completely destroy the crop at any stage. In some areas, the floodwater may be loaded with sediments that can cover the leaves and obstruct plant photosynthesis (McLean et al., 2002). Only minimal or no inputs are applied to deepwater rice crops due to the high uncertainty of harvests. Once mature, panicles are harvested, dried, and threshed. Farmers go by canoe or wait until the water recedes to harvest the panicles of floating rice (Murray, 2005). Tidal rice is cultivated during the rainy season in the coastal wetlands of East and West Africa. Tidal rice can tolerate submergence or flash floods to a great extent. Areas with saltwater intrusion from the sea require salinity‐ tolerant rice varieties. About 35‐ to 40‐day‐old rice seedlings raised in nurseries are transplanted in prepared main fields at 3 or more seedlings per hill (Gibba, 2003). In areas of poorly drained and acid sulfate soils, seedlings are transplanted on ridges to reduce Al and Fe toxicity (WARDA, 2002). It is recommended to apply 68‐22‐22 kg ha1 of N‐P‐K to mangrove rice in the Gambia (Gibba, 2003). Other practices are similar to those for rice in the rainfed wetland ecosystem. Nonirrigated WS rice yields range from 0 to 4.0 Mg ha1 depending on the season, location, and rice type. Floating rice yields are low, 1.0–2.5 Mg ha1. Yields of tidal rice vary widely and complete crop failure can occur in salt‐ aVected coastal wetlands. Although the mean rice yield is 1.0 Mg ha1 in mangrove swamps of the Gambia, improved rice varieties with adequate fertilization can yield 3.2–3.3 Mg ha1 in research plots (Gibba, 2003). WARDA has identified 341 promising rice varieties with target yields of 3 Mg ha1 or more in mangrove swamps (WARDA, 2002). d. Production Constraints. Major production problems in deepwater areas are submergence, salinity, acid sulfate soils, and Fe and Mn toxicity in coastal wetlands (Sahrawat, 2004a), and peat soils and cold injury at the seedling stage in high‐altitude areas (Balasubramanian et al., 1995). Farmers obtain low and extremely variable yields from nonirrigated WS rice crops due to the use of low‐yielding but adapted traditional varieties, the application of few or no inputs, and multiple environmental stresses—soil problems and unpredictable combinations of drought and flood. There is a lack of suitable plant types tolerant of submergence, salinity, acidity, Fe and Mn toxicity, and cold at the seedling stage (in highlands). 3.

Irrigated Wetland Rice Ecosystems

Irrigated rice is grown in bunded fields with assured irrigation for one or more crops per year. Usually, farmers try to maintain 0.05–0.1 m of water in rice fields. Irrigated rice areas are concentrated mostly in the humid,

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subhumid, semiarid, and high‐altitude tropics of the continent. Dams across rivers, diversion of water from rivers, or tube wells provide water for irrigation. We can distinguish three types of irrigated rice ecologies in SSA: the irrigated rice in the arid and semiarid Sahel, the irrigated rice in the humid forest and savanna zones (Defoer et al., 2002), and the irrigated rice of the tropical highlands (Balasubramanian et al., 1995). Irrigated wetland rice in the Sahel is akin to DS fully irrigated rice in Asia.

In the Sahel, solar radiation is high, water control is good, mechanization is widespread, pests are less prevalent, and input use is relatively high. Farmers are organized and infrastructure is well developed. About half of the irrigated rice area in the Sahel is direct seeded and the rest transplanted. Only 10% of the area is double cropped (Defoer et al., 2002). Per‐season rice yields are high, 5–8 Mg ha1 in research plots and 4–5 Mg ha1 in farmers’ fields vis‐a`‐vis potential yield of 8–12 Mg ha1 (Haefele and Wopereis, 2004). Mean rice yields in the OYce du Niger in Mali have increased from 2 Mg ha1 in 1977 to 6 Mg ha1 in 2002 (Defoer et al., 2002). However, in the DS, evapotranspiration is high and water consumption is considerably greater than in the WS. In addition, extreme temperatures limit rice yields in the WS and DS in some parts of the Sahel. Irrigated wetland rice in the humid forest and savanna zones is mostly transplanted. Irrigation schemes are small and located near main roads and towns. In the WS, rainfall is the main source of water for crop growth, and irrigation is used as a supplement during crop establishment and early crop growth periods as well as during mid‐season dry spells. Both the potential yields of 5–8 Mg ha1 per season and the actual yields of around 3 Mg ha1 per season are relatively lower for irrigated rice in the forest and savanna zones than in the Sahel (Defoer et al., 2002), mainly due to lower solar radiation, poorer water control, and higher pest incidence. In addition, iron toxicity is a constraint in irrigated wetland rice in the savanna and humid zones of West Africa, where tolerant varieties and nutrient management showed promising results on a long‐term basis (Sahrawat, 2004a; Sahrawat et al., 1996). Irrigated wetland rice in tropical highlands, in East and Central Africa and Madagascar, is grown on higher elevation marshlands or inland valley swamps at 700–900 m above msl (low altitude) and 900–>1600 m above msl (medium altitude). In well‐developed marshlands, rice land is terraced and fields are fairly level; small earth dams built across streams and small rivers at the higher level provide good water control and irrigate from 50 to >1000 ha. There are two seasons: the WS is from July to December and the DS from January to June. Water supply from rainfall and irrigation is good during the WS and erratic during the DS; flooding may be a

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problem in the WS in undeveloped marshlands. All wetland areas (100%) are planted to rice in the WS and only 60–70% of the area in the DS; in the remaining area, WS rice is rotated with wheat, maize, potato, or vegetables. Low temperature or cold injury often aVects the seedlings in the nursery or main field in June–July. As elsewhere in SSA, P is the most limiting nutrient to rice production in highlands. Achievable rice yields are high (7–9 Mg ha1) in higher elevation marshlands because of warm days (20–30 C) and cool nights (10–20 C) and longer duration (150–180 days). Rice yields decrease with increasing OM content: 7.3 Mg ha1 per season in soils with OM < 40‐g kg1 soil, 6.1 in soils with OM between 40‐ and 80‐g kg1 soil, and 4.7 in soils with OM > 80‐g kg1 soil (Balasubramanian et al., 1995). a. Irrigated Wetland Rice Area. Globally, the harvested irrigated rice area of 79 million ha produces more than 75% of the world’s rice output of 600 million Mg or more. In SSA, an estimated 19.8% of the cultivated rice area was irrigated (1.42 million ha) during 1995–2004 (Table V). Most of the WS rice areas are located in the rainforest and moist savanna zones of SSA. About 60% of the DS irrigated rice area of West and Central Africa is found in the Sahel, Sudan, and savanna zones (WARDA, 2002). High‐altitude irrigated rice is found in East and Central Africa and Madagascar. Countries with significant irrigated rice areas are Madagascar (621,764 ha), Nigeria (419,288 ha), Mali (80,366 ha), Guinea (49,550 ha), Senegal (39,600 ha), and Coˆte d’Ivoire (35,063 ha) (Table VI). b. Cropping Systems. Rice is the main crop during the rainy season and the land is left fallow for the rest of the year. The rice–rice‐fallow cropping system is practiced in double‐cropped irrigated wetlands of the Sahel (Defoer et al., 2002). Rice may be rotated with maize, soybean, or vegetables in inland valley bottoms of West Africa. In double‐cropped areas of tropical highlands (East and Central Africa and Madagascar), two rice crops are grown or WS rice is rotated with potato, wheat, soybean, or vegetables in the winter–spring season (Balasubramanian et al., 1995). c. Cultivation Practices and Yields. Rice fields are soaked with water, plowed, puddled, and leveled before crop establishment. Compost and animal manure, if applied, are incorporated into the soil during land preparation. Farmers use hand tools, animal‐drawn implements, or hand tractors to prepare the land. Large tractors are used only in large public and private sector irrigation schemes. Most of the WS irrigated rice is transplanted in

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rainforest and savanna zones and coastal plains, whereas 50% of the irrigated rice in the Sahel is direct seeded (Defoer et al., 2002). Seedlings are prepared in nurseries and 20‐ to 30‐day‐old seedlings are manually transplanted in main fields. Farmers weed their rice fields manually and apply some amount of fertilizers and pesticides, depending on availability and cost in local stores. Herbicide use to control weeds is minimal. At maturity, rice is harvested, dried, and threshed manually. In the Sahel of West Africa, mechanization is widespread, especially for land preparation and threshing (Defoer et al., 2002). With good water control and crop management, potentially irrigated DS rice yields in the semiarid Sahel zones can be as high as 8–12 Mg ha1 (Haefele and Wopereis, 2004). WARDA has identified 21 varieties with target yields higher than 5 Mg ha1 for irrigated rice in the Sahel and 426 varieties with yields higher than 4 Mg ha1 for irrigated lands in humid and subhumid forest zones (WARDA, 2002). However, in most of the irrigated areas of SSA, farmers obtain 2–5 Mg ha1 due to irregular irrigation, poor soil and crop management, and inadequate input use (Mie´zan and Sie, 1997). The gap between attainable and average farmers’ yields is 2–6 Mg ha1 in the Sahel, 2–5 Mg ha1 in the rainforest and savanna zones (Defoer et al., 2002), and 2–6 Mg ha1 in the tropical highlands (Balasubramanian et al., 1995).

d. Production Constraints. Many of the irrigated rice production constraints are similar to those in Asia: poor land preparation, leveling, and irrigation management; inadequate drainage leading to the development of salinity and alkalinity (inland basins and coastal wetlands); poor management of production inputs; yield instability due to weeds (in direct‐seeded rice), insect pests, and diseases; and deteriorating irrigation infrastructure, especially in large public irrigation schemes (Defoer et al., 2002). P deficiency in all soils, N and P deficiency in mineral and slightly organic hydromorphic soils, Fe toxicity at poorly drained sites, and high OM content in peat soils limit irrigated rice yields in the tropical highlands (Balasubramanian et al., 1995; Sahrawat, 2004a; Sahrawat et al., 1996). K or Si deficiency increases the susceptibility of rice crops to diseases. In addition to blast disease, Africa‐specific biotic constraints include AfRGM, RYMV, and glume discoloration. Stem borers, rats, and birds are other pests that attack rice in all ecosystems. Many large public sector irrigation projects have not been successful due to a combination of factors (Defoer et al., 2002). As a result, irrigated rice yields have declined from more than 7 Mg ha1 at the start of many irrigation projects to less than 3 Mg ha1 after a few years. Smaller farmer‐ managed irrigation schemes may be a viable alternative for sustainable irrigated rice production in many African countries. At the same time, rigorous

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socioeconomic studies exploring the causes for failure of large‐scale irrigation schemes are called for.

VI.

RICE PRODUCTION CONSTRAINTS IN SSA

Biophysical, management, human resource, and socioeconomic/policy constraints plague rice farming in SSA.

A. PHYSICAL, BIOLOGICAL, AND MANAGEMENT CONSTRAINTS Physical, biological, and management constraints vary with rice ecosystems as discussed above in Section V.

B. HUMAN RESOURCE CONSTRAINTS Particularly serious is the lack of researchers. As is pointed out by Evenson and Golin (2003), the ratio of researchers to extension workers is much lower in SSA than in Asia. This is truly a serious problem because the lack of profitable technology, but not the lack of extending it, is the most basic constraint to improving farming eYciency in SSA. Once new profitable technologies are developed, demand for extension services will increase. In such a situation, it is expected that capacity enhancement programs for extension workers, which will have high payoVs, will be undertaken. The lack of education among rice farmers is another major constraint, as better‐educated farmers are more willing to adopt new technologies (Schultz, 1975). According to the Green Revolution experience in Asia, however, the role of education becomes less important over time as the new technology is widely adopted (David and Otsuka, 1994). Given the substitutability between farmers’ education and extension services, and the lack of education among farmers in SSA, strengthening the extension system is likely to be an appropriate strategy once new technologies become available. In the longer run, the continuous development of new technologies will attract educated farmers to engage in scientific rice farming and stimulate investments in schooling of children. Other human resource‐related constraints are as follows: Weak or nonexistent research‐extension‐farmer linkage Poor or no farmer organizations Lack of public–private partnerships

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Reduced labor availability due to poor nutrition and/or diseases such as

AIDS, cholera, malaria, bilharzia, and so on.

C.

SOCIOECONOMIC

AND

POLICY CONSTRAINTS

In addition to biophysical and human resource constraints, rice production in SSA is aVected by socioeconomic and policy constraints: Unfavorable input and output pricing policies at the national level. Low

output prices vis‐a`‐vis high and rising input prices reduce profit and the competitiveness of smallholder farms in local, regional, and global markets. Limited access to credit, inputs (seed, fertilizers, pesticides, implements, and so on), markets, and market information. Poor rural infrastructure and transportation. This unfavorable price structure reflects the ineYcient marketing systems in SSA. The establishment of eYcient marketing systems requires trust between local traders and farmers and between local and urban traders, because dishonest behavior, such as cheating on product quality and late delivery, can easily occur in any transaction (Hayami and Kikuchi, 2000). To prevent such behavior, trust must be developed through long‐term and repeated transactions. Prerequisites for such development are (1) the improvement of rural infrastructure and transportation systems and (2) the availability of fertilizer‐responsive‐improved varieties and eYcient technologies that enhance the profitability of long‐term transactions between farmers and traders. Experience shows that socioeconomic institutions are not rigid, but are subject to change as new profitable opportunities arise not only in Asia (Hayami and Kikuchi, 1982; Hayami and Ruttan, 1985) but also in SSA (Otsuka and Place, 2001). How such development may take place and whether any major obstacles to change exist in SSA need to be analyzed through collaborative research between social scientists and researchers engaged in the development of new rice technologies.

VII. RICE RESEARCH AND TECHNOLOGY DEVELOPMENT DURING THE PAST 20 YEARS The two leading international rice R&D institutions—International Rice Research Institute (IRRI) and Africa Rice Center (WARDA)—complement each other in developing the rice sector in SSA and bringing concrete benefits to its rice farmers and consumers. During the past two decades,

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both centers have generated many rice research outputs and technologies through individual and collaborative research with national agricultural research institutes (NARIs). These research findings and technologies are discussed briefly in this section.

A.

RICE GERMPLASM, BREEDING, 1.

AND

VARIETY DEVELOPMENT

Plant Types and Traditional Rice Varieties of SSA

The genus Oryza has two cultivated species, O. glaberrima Steud (African rice) and O. sativa (L.) (Asian rice), and 21 wild taxa of Asian and African origin. Of the 21 wild rice species, seven originated in SSA and the rest in Asia. The seven African wild rice species and their characteristics are given in Table VII. These wild species carry genes for specific traits, for example, resistance to biotic and abiotic stresses, and these genes are valuable in interspecific crossing (Khush, 1997). O. glaberrima is the African indigenous species, propagated from its original center, the upper‐middle delta of the Niger River, and extended toward the Senegal, Gambia, Casamance, and Sokoto basins (Carpenter, 1978; Murray, 2005). It has been cultivated for 3500 years (Dembele, 1995; Porte`res, 1956). O. glaberrima has no subspecies. It is adapted to African environments but prone to lodging and grain shattering when compared with O. sativa (Sie, 1991). It possesses early plant vigor and resistance to African stresses such as drought, blast, RYMV, nematodes, and insects (Adeyemi and Vodouhe, 1996). These qualities are valuable in interspecific crossing. O. sativa has two subspecies—indica and japonica—with a continuous array of intermediates. The many agroecotypes of O. sativa are adapted to various growing conditions and improved varieties derived from them are highly productive. The japonica is adapted to rainfed drylands and indica to the aquatic ecology of SSA. But they are susceptible to many abiotic and biotic stresses in SSA (Sie, 1991). O. glaberrima is now being replaced by the Asian rice O. sativa introduced by European traders into SSA around the sixteenth century (Buddenhagen, 1986; Porte`res, 1970). Some farmers grow both Asian and African rice side by side to meet their varied needs and to tackle adverse conditions. They like the African rice for its fast early growth that can suppress weeds, its shorter duration, its resistance to diseases, and the nutty flavor of its grains (Nyanteng et al., 1986). Varietal classification is done based on the characteristics of the two cultivated species and cultural practices in the field—time of inundation,

88 Table VII African Wild and Cultivated Rice Species and Their Origin, Distribution, Plant Type, Grain Features, and Valuable Genes for Interspecific Crossing

I. Wild species O. barthii (O. breviculata) A. Chev. and Roehr. (ancestor of O. glaberrima) O. longistaminata A. Chev. and Roehr. (possible ancestor of O. barthii) O. punctata Ktoschy ex Steud.

O. brachyantha Chev. and Roehr.

Origin and distribution

Plant type and grain features

Valuable donor genes

AgAg

West African Savanna and Sahel (swamps and waterholes)

An annual grass; self‐fertile; long grains w/awns, shattering; dormancy

Resistance to GLH, bacterial blight; drought avoidance

Khush (1997); Vaughan (1994)

24

AgAg

Tropical Africa and Madagascar

A tall, erect perennial w/rhizomes; cross‐ pollinated; long thin grains w/awns; shedding

Resistance to bacterial blight; drought avoidance

Besanc¸on et al. (1978); Khush (1997); Vaughan (1994)

24 48

BB BBCC

Both annual and perennial; long, narrow grains

Resistance to BPH, leafhopper

Katayama (1990); Khush (1997); Vaughan (1994)

24

FF

East and West Africa; Madagascar (forest and waterholes) West, Central and South‐eastern Africa

A short, slender annual; small and very narrow grains w/long awns

Resistance to yellow SB, leaf folder, whorl maggot, of tolerance to lateritic soil

Khush (1997); Vaughan (1994)

2n

Genome

24

References

V. BALASUBRAMANIAN ET AL.

Species

24

AgAg

West Africa

A perennial weedy species

–

Bradenas and Chang (1966) Vaughan (1994)

O. eichingeri

24 48

CC BBCC

Short, sturdy annual and perennial; short grains

–

O. schwein furthianaa

48

BBCC

East, Central and S.E. Africa; Sri Lanka Tropical Africa

A perennial

–

II. Cultivated species O. glaberrima Steud.

24

AgAg

West Africa

Cultigen; fast early growth, weed suppression

Bradenas and Chang (1966); Khush (1997); Vaughan (1994)

O. sativa (L.)

24

AA

Asia

An annual; dryland erect photoinsensitive; and floating photosensitive; no second or third branching in panicles; short grains, shedding Indica and japonica subspecies w/ intermediates; erect to floating; traditional and improved

Cultigen; forked branches in panicle, no grain shedding

Khush (1997); Vaughan (1994)

INCREASING RICE PRODUCTION IN SUB‐SAHARAN AFRICA

O. stapfii Roschev.

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flood duration, maximum water depth, and level of soil fertility. On the basis of these criteria, five varietal groups are recognized in SSA: 1. ‘‘Rainfed’’ varieties called ‘‘mountain rice’’ or ‘‘dryland rice’’ are often cultivated in watershed areas and along forest galleries and in upper parts of inland valleys where flooding is not common. These plants are low in tillering and suited to direct seeding. 2. ‘‘Early‐duration erect varieties adapted to a submerged environment’’ are often found on valley fringes with hydromorphic soils having variable moisture regimes. 3. ‘‘Season‐erect varieties’’ are often planted on lower parts of the toposequence than the previous ones and are subject to flooding (0.5‐ to 0.8‐m water depth). They are suitable for rainfed wetlands. 4. ‘‘Late‐duration erect varieties’’ are mostly adapted to river floodplains and deepwater areas. 5. ‘‘Floating’’ varieties capable of surviving floods remain above water level even at significant water depths (1 to >3 m). O. glaberrima is predominant in this category. 2.

Glaberrima sativa Crosses and the Development of NERICA Varieties for African Drylands

In 1991, Dr. Monty Jones of WARDA led a team of scientists in a new breeding eVort to unlock and combine the genetic potential of Asian and African rice. Key to the eVort was WARDA’s rice Gene Bank with 16,000 rice varieties preserved in cold storage, and duplicated at the International Institute for Tropical Agriculture (IITA) in Nigeria and at IRRI in the Philippines. Among the preserved varieties are 1500 O. glaberrima lines. Jones and his group used molecular biology to overcome sterility, the main problem in interspecific crossing, and to speed up the breeding cycle. They made ‘‘wide crosses’’ of the African and Asian rice, then removed the fertilized embryos by embryo rescue and grew them in artificial media. They then backcrossed the progeny twice to the Asian parent to recombine the genetic backgrounds of the two distinctly diVerent species. Backcrossing allowed the introgression (merging) of useful genes such as the wide, droopy leaves from the rugged O. glaberrima into the more productive O. sativa subspecies japonica background. Anther culture helped breeders ‘‘fix’’ progeny lines rapidly and retain recombinant lines—with the combined traits of both the African and Asian parents (Fig. 7). With conventional breeding where the progeny of a cross segregates into diVerent plant types, it takes 5–7 generations to isolate, purify, and select a line with a desired combination of genetic traits. For dryland rice, this could mean 5–7 years because usually only one crop can be grown per year. With anther culture, a line can be

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Hybridization scheme for the production of NERICAs Oryza glaberrima African rice

Oryza sativa Asian rice

× F1

×

O. sativa

BC1F1

×

O. sativa

BC2F1 Pedigree selection

Anther culture Ho

Fixed lines (new plant type) NERICA

BC2F6

BC2F1

Figure 7 Hybridization scheme for producing the new rice (NERICA) varieties for SSA drylands.

selected after only one generation, and a new variety developed in 18–24 months. By the mid‐1990s, WARDA scientists were testing the new rice for Africa (NERICA) in rainfed drylands (WARDA, 2001–2002). Traits of dryland NERICA varieties: The genetic diVerences of the two distant species crossed gave NERICA varieties high levels of heterosis or hybrid vigor for faster growth, higher yield, and more resistance to stresses than either parent. The NERICA varieties have raised the ‘‘yield ceiling’’ of dryland rice by 50%, from the current level of 4 Mg ha1, due to longer panicles with forked branches bearing 2–3 grains each and more grains per panicle (Table VIII). The new varieties are taller than O. glaberrima, which makes harvesting easier—especially if the woman farmer has a baby strapped to her back. The shorter duration (90–100 days) of NERICA varieties allows farmers to grow two crops during one rainy season—one rice crop and a post‐rice dryland crop such as grain or fallow legumes that can smother weeds and add up to 60 kg of biologically fixed N per hectare. Each hectare of well‐ managed rice–legume rotation could save 4 ha of fallow land from clearing. Projections on the spread of NERICA varieties and rice‐fallow legume rotations in West Africa indicate that by 2010 the total area of land saved will be about 15,000 ha (WARDA, 2001). O. glaberrima is highly tolerant of drought. When faced with drought, the thin leaves ‘‘roll’’ quickly to retain water and the thin roots grow deep into the soil to explore water. The total length per gram of O. glaberrima roots can reach about 150 m in contrast to 100 m for O. sativa roots. When the rains come after drought, the O. glaberrima recovers faster because the replacement of the thin leaves and roots requires less water and nutrients.

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Table VIII Improved Traits of the O. glaberrima O. sativa Crosses (NERICAa) Traits

O. glaberrima parent

O. sativa parent

O. glaberrima O. sativa crosses (NERICA lines)

Short

Short

Taller than both parents

Leaves

Droopy, wide

Erect

Stems

Strong, sturdy

Weak, thin

Droopy, wide lower leaves at early vegetative growth and erect upper leaves at reproductive phase Strong

Tillering

Low

High

High

Panicle type

Medium long with no forked branches Single Low (75–180) High

Long with forked branches 3–4 High (100–250) No

Longer than either parent with forked branches 3–4 Very high (up to 400) No

150–170

120–140

90–100

Thin leaves and long thin roots (150 m g1), thin leaves and roots recover fast with rains after drought

Medium thick leaves, short roots (100 m g1), slow recovery after drought

Thin leaves and long thin roots, thin leaves and roots recover fast with rains after drought

Grains per branch Grains per panicle Grain shattering at maturity Duration (days) Drought tolerance traits

a

NERICA, New Rice for Africa.

Facilitates harvesting of panicles without bending Early weed suppression and higher photosynthesis at reproductive phase > higher grain yield Help bear heavy panicles, no lodging More productive tillers, higher yield Higher yield

No loss of grain at harvest Higher yield per day, allows double cropping Good drought tolerance and avoidance, fast recovery with rains after drought

V. BALASUBRAMANIAN ET AL.

Plant height

Inherited advantages of NERICA lines

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The NERICAs have inherited the thin leaves and roots of O. glaberrima (WARDA, 2001–2002, 2002–2003).

3.

Breeding of New NERICA Varieties for African Rainfed Wetlands

Rainfed wetland rice in Africa suVers from variable rainfall, unpredictable drought and flooding, AfRGM, RYMV, and blast. Most of the traditional wetland rice varieties grown in the region have a narrow genetic base, which leads to their vulnerability to drought, diseases, and pests. Some diseases, such as RYMV, are spreading fast in the region because of the predominant cultivation of susceptible rice varieties. Therefore, Dr. Sie of WARDA and his partners crossed specific RYMV‐resistant African rice varieties with popular—but susceptible—Asian rice (O. sativa subspecies indica) varieties. As can be envisaged, the initial problem was hybrid sterility (infertile oVspring of the crosses) because the two rice species have evolved separately over millennia and are so diVerent that often attempts to cross them do not lead to reliable variety development (WARDA, 2003–2004, 2004–2005). The sterility problem is greater when we cross the African rice with indica than with japonica. The sterility blockage was overcome by backcrossing (crossing the hybrid with an O. sativa parent) to restore fertility. Some of the progeny combined the best features of both parents: the droopy leaves and vigorous early growth (associated with weed competitiveness) typical of the African rice and the high number of spikelets (indicating productivity) of the Asian rice. A major scientific milestone was achieved when the screening for resistance to RYMV under artificial infestation showed that the crosses had successfully transferred resistance to RYMV into some of the progeny. A new plant type with high yield potential is now available for wetlands, endowed with resistance to local stresses, particularly to the dreaded RYMV. The progeny of O. glaberrima and O. sativa subspecies indica are better adapted to rainfed and irrigated wetland conditions, while those of O. glaberrima and O. sativa subspecies japonica are more suited to rainfed dryland conditions in SSA (WARDA, 2003–2004, 2004–2005). The shuttle‐breeding approach used between WARDA breeders and national programs facilitated the fast exchange and evaluation of breeding lines under diVerent conditions, accelerated the selection process and increased its eYciency, and helped achieve wide adaptability of new plant types. For example, in Burkina Faso, about 600 new plant‐type lines were tested in the wetlands of the Banfora research station for 4 years (2000–2003) and the 20 most promising lines were selected based on yield and resistance to stresses, especially to RYMV. Lines of the new plant type were also evaluated in other important rice‐growing countries of West Africa—Mali, Burkina Faso, Togo, and Senegal—and more than 70 promising lines were selected.

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In addition, farmers were involved in the early varietal selection process through farmer participatory variety selection (PVS). The PVS approach helped farmers choose varieties that meet their needs and breeders obtain feedback from farmers regarding their preferences for plant type and grain characteristics and speed up the fine‐tuning, adoption, and dissemination of new varieties. The PVS exercise showed clearly that men farmers gave importance to short growth duration and plant height, whereas women preferred traits such as good emergence, seedling vigor, and droopy leaves that indicate weed competitiveness, since they are mostly involved in sowing and weeding operations. The three most preferred new plant‐type lines are WAS 122‐IDSA‐ 1‐WAS‐B‐FKR‐B‐1, WAS 122‐IDSA‐1‐WAS‐2‐FKR‐B‐1, and WAS 122‐ IDSA‐1‐WAS‐6–1‐FKR‐B‐1. They have a yield potential of 6–7 Mg ha1, good tillering ability, growth duration of 120 days, and an acceptable plant height; all three varieties showed good resistance to major wetland stresses and also responded well to N application. Four new wetland NERICA varieties, now oYcially known as the ‘‘Wetland NERICAs,’’ have been released in Burkina Faso, two in Mali, and three in the Gambia (WARDA, 2003–2004, 2004–2005). The wetland NERICA varieties oVer a powerful new weapon to rice farmers to manage their complex wetland rice stresses. However, to be most eVective, they should be used as part of the integrated crop management (ICM) approach developed by WARDA. Work continues with genetic engineering and molecular tools to identify and incorporate traits for various abiotic stresses into elite rice germplasm and to develop suitable wetland varieties through both interspecific crossing (crosses between the two cultivated species of rice) and intraspecific crossing (crosses within the species, i.e., between O. sativa varieties). Table IX provides a summary of ongoing conventional and molecular‐assisted breeding eVorts at IRRI and WARDA to develop rice varieties suitable for diYcult rice environments in SSA (Gregorio et al., 2006).

4.

Development of Improved Irrigated Rice Varieties

Since 1971, IITA has been involved in varietal improvement research for irrigated rice in Africa and in the national and international varietal‐testing network in the region. The strategy for irrigated rice improvement was to incorporate resistance to or tolerance of Africa‐specific stresses such as RYMV, blast, AfRGM, and so on, into promising introductions from Asia and Latin America (Buddenhagen, 1986). Through this program, IITA has identified or developed high‐yielding rice varieties suitable for irrigated systems: ITA 212, ITA 222, and ITA 306, with a duration of 120–130 days, a

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Table IX Priority for Genetic Engineering and Conventional Breeding Approaches for Incorporating Resistance to/Tolerance of Abiotic Stresses in Rice Priority Traits for abiotic stresses

1

2 a

3 b

P deficiency Zn deficiency

MAS MAS

CON CON

– –

Salinity

MAS

CON

GMRc

Drought

MAS

CON

GMR

Submergence

MAS

CON

GMR

Fe toxicity Al toxicity

MAS MAS

CON CON

– –

Cold tolerance Elongation ability

MAS MAS

CON CON

– –

Comments G E interactions favor use of MAS Genes introgressed from wild species into O. sativa and MAS under progress MAS allows pyramiding genes of QTLs for diVerent mechanisms of tolerance MAS and GMR under development through functional genomics MAS in full implementation with the identified Sub1A gene G E interactions favor MAS Genes introgressed from wild species into O. sativa and MAS under progress G E interactions favor MAS Genes introgressed from wild species into O. sativa and MAS under progress

a

MAS, DNA marker‐aided selection (includes use of linked markers, and candidate genes or identified genes). b CON, conventional breeding. c GMR, genetically modified rice. Adapted from Gregorio et al. (2006).

plant height of 0.97–1.06 m, and a mean yield of 4.7–5.4 Mg ha1 in national trials in Nigeria and Cameroon; two Fe‐toxicity‐tolerant varieties, ITA 247 and ITA 249; and two cold‐tolerant varieties, B2161‐C‐MR‐51‐1‐3‐1 and IR7167‐33‐2‐3 for high‐altitude areas. Donors identified for incorporating resistance to African stresses were Moroberekan, LAC 23, ITA 235, and CT 19 for RYMV; Cisadane and Eswarakora for AfRGM; and ITA 121 and DJ 12‐539‐2 for stalk‐eyed fly (Masajo et al., 1986). The NARES introduced a large number of irrigated rice varieties from abroad—Japan, United States, Thailand, China, Portugal, Spain, Egypt, and Madagascar. However, most of the current irrigated rice varieties were introductions of the 1970s through WARDA–INGER (International Network for Genetic Evaluation of Rice) coordinated regional trials in West Africa (WARDA, 1996). The local irrigated rice‐breeding program initiated in the mid‐1970 exploited both African and exotic germplasm to develop improved irrigated rice varieties (Sie, 1994).

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5.

Germplasm Exchange by INGER‐Africa

In 1975, IRRI launched the global International Rice Testing Program (IRTP) for the systematic collection, distribution, and testing of rice genetic materials. Later, it became INGER, with the objectives of promoting global exchange, evaluation, and use of improved breeding materials originating from sources worldwide. INGER‐Africa was part of the global program until it was transferred to WARDA in April 1997. INGER‐Africa had 16 types of regional rice evaluation trials/nurseries targeted for dryland, irrigated, rainfed wetland, and mangrove ecosystems and biological stresses of blast and RYMV. From 1985 to 1996, INGER‐Africa distributed 3726 nursery sets to African countries. Only a few rice varieties adapted to local conditions were selected by national breeders from the INGER‐Africa nurseries due to the mismatch between national capacity and needs and the supply from the program. A new germplasm exchange mechanism began in 1991 and was formalized in 1994 at WARDA to modify certain operational aspects of INGER‐Africa to make it more eYcient and responsive to national needs in SSA. With the new approach, nurseries are designed to (1) fit NARES’ needs and avoid overloading their capacity, (2) provide genetic diversity and variability for key rice ecosystems, and (3) target a supply of valuable germplasm. In addition, the new mechanism oVers NARES: Improved germplasm from a wide range of sources for national breeding

and direct selection of varieties—nursery entries come from several countries or institutions in Africa (73%), Asia (19%), and Latin America, Europe, the United States, and other countries (8%) A mechanism to screen their own genetic materials for resistance to or tolerance of specific stresses at reliable hot spot locations in the region The means to test the agronomic stability and adaptability of their elite varieties in regional multilocation trials The means to handle segregating populations—F3 populations nominated by breeders are grown and harvested in bulk for distribution to NARES on request for in situ selection and advancement nationally. With the creation of the WARDA research task forces in 1991, members of task forces who are also members of the new INGER‐Africa meet every year to report on their results over the previous year and plan their activities for the following year. This arrangement has increased interactions among scientists and improved the rate of return of trial results. The NARES’ share of varieties in regional nurseries increased from <10% in 1985 to about 60% in 2000. Although most of these rice germplasm materials were bred or selected in Africa, their parents came from wide sources: India, Sri Lanka, Bangladesh,

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Mozambique, Indonesia, Philippines, Thailand, Taiwan, China, Colombia, Argentina, Brazil, Zaire, Sierra Leone, Coˆte d’Ivoire, Nigeria, Guinea, Mali, Senegal, the Gambia, and international research centers such as IRRI, IITA, WARDA, and the International Center for Tropical Agriculture (CIAT). Thus, despite limited regional resources invested annually in varietal improvement, 197 improved varieties had been released up to 1999, with more than 122 during the next 5 years (2000–2004). These results involve only 7 out of 17 WARDA member countries and indicate the eYciency and eVectiveness of the new approach to germplasm exchange and evaluation in SSA. In addition to germplasm exchange, INGER‐Africa activities include (degree, group, and in‐country) training of NARES scientists and technicians in the region.

B. RICE SEED PRODUCTION AND DISTRIBUTION SERVICES One of the principal strategies for improving food security in SSA is to strengthen the seed supply sector (FAO, 1997). Unlike fertilizers and pesticides, which are often available and used, particularly on cash crops, good‐ quality seeds are rarely used by farmers in SSA. Guaranteeing farmers access to good‐quality seed can be achieved only if there is a viable seed supply system to multiply and distribute seeds of improved varieties and if mechanisms to assist farmers in emergency situations have been established. As long as the seed varieties oVered are well adapted to small‐farm environments and low‐input crop management practices, other inputs such as fertilizers are less critical compared to the benefits derived from improved seed.

1.

Status of Existing Seed Sector Services and Seed Research

To design realistic strategies for the future development of the seed supply sector in SSA, it is imperative to assess in‐depth the existing seed sector. There are formal and informal seed sectors in all countries. Formal seed supply systems include public sector institutions, such as parastatal (quasigovernment) seed agencies providing seed certification and quality control, and the private sector. The formal seed production and supply programs are organized mostly by the public sector and commonly assisted by donor agencies. Private companies are also involved in the formal seed sector, especially in Ethiopia, Madagascar, Malawi, Mozambique, Nigeria, South Africa, Zambia, and Zimbabwe. However, despite these eVorts, formal seed supply systems currently meet not more than 5–10% of the seed needs of farmers in the region.

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Most formal seed production projects in SSA failed because of The use of the criteria and seed standards of the seed industry of developed

countries, which are diYcult to implement and sustain under developing country conditions The lack of follow‐up of donor‐funded programs to ensure continuity of activities after a program or project ended The concentration of donor‐funded projects exclusively on the formal seed sector, ignoring the already well‐established informal sector The concentration of almost all projects on major crops of commercial value, neglecting local crops of economic and social importance to small farmers. Informal seed supply systems are composed of indigenous strategies to improve the quality and quantity of seed used by farmers. About 90–95% of the seed production in almost all countries of SSA is still in the informal seed supply system. Seed research: The average investment by SSA countries in agricultural research is estimated to be less than 3% of gross domestic product (GDP). Thus, the primary seed research activities in most countries involve variety development and testing, variety release and registration, variety maintenance, and breeder seed production. Since variety development is time consuming and costly, many governments in the region tend to prioritize variety development research alone according to urgent national needs. As a result, research strategies are directed to a few cereals and staple food crops only.

2.

National Seed Laws and Variety Release Process

Only 25% of the African countries have passed a seed act that stipulates specific seed regulations that must be satisfied. The remaining 75% of the countries in SSA do not have legislation governing the production, distribution, and sale of seeds. In addition, even in most of those countries where a seed act has been passed, putting the various laws and policies into practice has been impeded by inadequate enforcement mechanisms and a lack of logistical, financial, and human resources. In addition, the variety release process is weak in many countries, and it may take 10 years to release and register a new variety in some countries. The documentation of new varieties with their origin and agronomic and technical characteristics is again poor due to the lack of trained staV and resources. The breeder holds an exclusive right to multiplication and distribution of seeds. There is no national strategy with a good awareness campaign to encourage the production, distribution, and use of good‐quality seeds widely.

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C.

99

CROP ESTABLISHMENT

Timely crop establishment is critical for good plant growth and high yields. In irrigated DS rice areas, optimum planting time can be maintained by a regular water supply. In rainfed drylands and wetlands, the arrival of rains decides the time of planting. Rice is mostly transplanted in irrigated wetlands and some rainfed wetlands. Half of the irrigated rice in the Sahel is direct seeded. In transplanted rice areas of West and Central Africa, research data recommend not to transplant in August or from November to December to avoid low‐temperature stress (WARDA, 2002). Good nursery management and transplanting of young and vigorous seedlings at optimum spacing will increase irrigated rice yields (Balasubramanian et al., 2005). In mangrove swamps, it is recommended to plant rice on ridges to improve drainage and to reduce Al and Fe toxicity. Farmers in Africa prefer direct seeding if suitable methods are available. In addition to seeding techniques, research must develop technologies to prepare smooth and level seedbeds for direct sowing and control weeds eVectively in direct‐seeded rice (Buddenhagen, 1986; Defoer et al., 2002). In rainfed drylands and wetlands, rice seeds are broadcast, drilled, or dibbled in prepared dry‐to‐moist soil. All three methods are equally eVective when the optimum seed rate of 50–80 kg ha1 is used (WARDA, 2002). In drilled and dibbled rice fields, eVective weed control is possible with interrow cultivation.

D. NUTRIENT MANAGEMENT Unlike dryland soils, which are mostly acidic, infertile, and highly prone to erosion, it is easier to keep soils fertile and stable in wetlands. Naturally, water and nutrients accumulate in wetlands; in addition, N is made available through microbial N fixation, P through increased solubility under reduced conditions, and K and other bases are supplied by rain and irrigation water. Of the 16 known elements required for plant growth, N, P, K, S, and Zn are important for rice; Si is becoming deficient in intensively cultivated rice areas of Asia (e.g., Sri Lanka). Other elements taken up by rice crops are Ca, Mg, Fe, and Mn. The importance of boron (B) to flooded rice is not established. To produce 1 Mg ha1 of grain yield, rice crops remove on average 19.1 kg ha1 of N, 2.8 of P, 23.6 of K, 3.9 of Ca, 5.2 of Mg, 0.50 of Fe, 0.55 of Mn, and 0.03 of Zn. Nutrient uptake by irrigated and dryland rice is similar for N, Mn, and Zn; irrigated rice removes more P, K, Ca, and Mg than dryland rice per megagram per hectare of grain yield and the reverse is true for Fe only (Table X) (Sahrawat, 2000). Dryland rice seems to be more eYcient in using P for grain production. In this study, the

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Table X Nutrients Removed (in Grain þ Straw) by Irrigated Wetland and Rainfed Dryland Rice Crops per Megagram per Hectare of Grain Yield in Coˆte d’Ivoire, West Africa, 1994 and 1996 Nutrients removed Nutrient N (kg ha1) P (kg ha1) K (kg ha1) Ca (kg ha1) Mg (kg ha1) Fe (g ha1) Mn (g ha1) Zn (g ha1)

Irrigated rice 18.9 3.4 28.7 4.8 6.8 384 561 30

a

Nutrient element harvest index (%) b

Dryland rice

Irrigated ricea

Dryland riceb

19.4 2.1 18.5 3.0 3.6 616 541 25

57 67 9 17 28 48 21 50

60 71 11 15 30 21 8 38

Irrigated rice: Mbe site; Ultisol; cv Bouake 189; 1996 DS; Yield: 6.77 Mg ha1. Dryland rice; Man site, Alfisol; cv WAB 56–50; 1994 WS; Yield: 3.14 Mg ha1. Data derived from Sahrawat (2000). a b

nutrient element harvest index [amount in grain/amount in (grain þ straw)] was the highest for P, followed by N and Zn, indicating the presence of a higher proportion of these three elements in the grain. About 90% of K, 85% of Mn, 84% of Ca, 71% of Mg, and 65% of Fe were retained in the rice straw (Sahrawat, 2000). Thus, the return of straw to rice fields will help reduce the depletion of K, Ca, and Mg from rice soils. For any system to be sustainable, the export of nutrients from the system through harvests must be balanced by nutrient additions. Integrated nutrient management considers all sources of nutrients interacting in a system. Essentially, it involves three principles (Smaling et al., 2002): Add new nutrients to the system—composts and manures from outside the

field, fertilizers, N fixation in wetland rice fields, and biological N fixation by legumes Save nutrients from being lost from the system—soil erosion control, return of crop residues, planting deep‐rooted crops to reduce leaching losses Recycle nutrients within the system to maximize productivity and nutrient‐use eYciency. Using the three principles, several nutrient management approaches have been developed: ecosystem‐based and toposequence‐based nutrient management developed by WARDA, Benin; site‐specific nutrient management (SSNM) developed by IRRI, Philippines; integrated soil fertility management (ISFM) developed by the International Fertilizer Development Center (IFDC)‐Africa, Togo; and integrated plant nutrient systems (IPNS)

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developed by the FAO, United Nations, Rome, Italy. We shall briefly look at the first three approaches for rice in SSA.

1.

Ecosystem‐Based Nutrient Management

Dryland rice ecosystem: Dryland rice soils (Oxisols and Ultisols) are invariably acidic, infertile, and highly P‐fixing (Sahrawat et al., 2003). The application of rock‐phosphate (rock‐P) or other P fertilizers is critical before crops respond to N and other nutrients. Finely ground and moderately to highly reactive rock‐P can be applied to acid dryland soils (Diamond, 1985). Rock‐P applied to legume fallows benefits the following rice crops more than its direct application to rice crops (Somado et al., 2003). Research recommends an application of magnesium limestone at 150–200 kg ha1 for slightly acidic soils and 2–3 Mg ha1 for highly acidic soils before fertilizers are applied to rice crops. Recommended fertilizer rates for dryland rice range from 40‐36‐64 to 90‐27‐36 kg ha1 of N‐P‐K (WARDA, 2002). P and K are applied at planting, while N is to be applied in 2–3 splits at basal, mid‐tillering, and panicle initiation (PI) stages. The field must be weeded before applying N fertilizers. However, farmers apply little or no fertilizer and thus mine the soil continuously. At least, they should be encouraged to apply available crop residues, composts, and animal manures (Erenstein, 2003); improve short‐season fallows by planting legume cover crops (Akanvou et al., 2001a; Carsky et al., 2001; Somado et al., 2003); rotate or intercrop rice with legumes (Akanvou et al., 2001b); and mulch soils with crop residues to preserve soil fertility (Erenstein, 2003). Rainfed wetland ecosystem: P is deficient in most wetland (Oxisols, Ultisols, Vertisols, and certain Inceptisols) and acid sulfate soils (Sulfaquepts, Sulfaquents). P availability is controlled by P sorption–desorption kinetics and solubility rates of various Al, Fe, and Ca phosphates in soils (Diamond, 1985). P adsorption by soils is related to their mineralogy: quartz ¼ Al‐free OM < 2:1 clays < 1:1 clays < crystalline Fe‐ and Al‐oxides < amorphous Al‐ and Fe‐oxides (Juo and Fox, 1977). When the soil is submerged, P availability increases from ferric and aluminum phosphates due to increasing pH in acid to neutral soils and from calcium phosphates due to decreasing pH in basic calcareous and alkaline soils (Ponnamperuma, 1972). The incorporation of P fertilizers during land preparation or surface application until 28 days after transplanting or dipping of seedling roots into a soil‐P slurry before transplanting is a good practice for P application to flooded rice crops. In degraded wetland soils with high P‐fixing capacity, P should be applied in two splits—at planting and PI (WARDA, 2002). Recommended N rates for wetland rice range from 40 to 90 kg ha1, applied in 2–3 splits (at planting, mid‐tillering, and PI). For rice crops in

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mangrove swamps and tidal wetlands, 40–80 kg ha1 of N in 3–4 splits and 40 kg ha1 of P as basal are recommended for they are the most limiting nutrients in this ecosystem (WARDA, 2002). Rice straw incorporation into soils may maintain soil K at low‐yield levels; but K fertilizer needs to be applied along with N and P to get high rice yields (Kyuma et al., 1986). In addition to NPK, plants require low amounts of S and micronutrients such as Zn, Fe, and so on. Soil submergence increases the availability of Fe, Mn, and Mo, but decreases that of S, Zn, and Cu; it also reduces Fe and Mn toxicity in acid soils, including acid sulfate soils (Neue and Mamaril, 1985). In addition, the slow decomposition of OM accumulated in flooded soils (Sahrawat, 2004b) may supply some of the micronutrients to rice plants. Thus, the need for the application of S and micronutrients to rice crops must be based on the knowledge of their established deficiency in soils. Nutrient forms in soils and sources for application, critical values for their deficiency and toxicity, and application methods are summarized in Table XI. Irrigated rice ecosystem: Inadequate nutrient application is among the key constraints to achieving high yields in irrigated rice where fertilizer use is most profitable (Haefele et al., 2000, 2001, 2002). Although rice varieties and production technologies, including fertilizer rates, were introduced from Asia, they were gradually adapted and modified to the local biophysical and economic conditions (Haefele and Wopereis, 2004). Farmers’ N rates varied from 37 to 251 kg ha1, P from 9 to 66 kg ha1, and K from 0 to 7 kg ha1 in Senegal and Mauritania, which represent the irrigated rice systems in the Sahel zone of West Africa. Average grain yields were 4.1–5.6 Mg ha1 vis‐a`‐vis the simulated potential yields of 8.0–10.0 Mg ha1 for the Senegal River Valley (Table XII). An estimated 46–70% of the applied N is lost from the soil‐plant‐water system due to variable rates and poor timing of N application (Haefele and Wopereis, 2004). Haefele and Wopereis (2004) improved irrigated rice fertilizer recommendations for five AEZs of the Senegal River Valley, based on a simple three‐ quadrant model, farmer surveys, on‐station experiments, and simulated potential yield for the WS and DS. The blanket recommendation of 120‐26‐50 kg ha1 of N‐P‐K for all five AEZs was changed to 161‐20‐0 kg ha1 (AEZ I, II) and 135‐20‐0 kg ha1 (AEZ III, IV, V) for the WS. For the DS, they recommended 187‐25‐0 kg ha1 (AEZ I), 161‐20‐0 kg ha1 (AEZ II), and 135‐20‐0 kg ha1 (AEZ III); no DS recommendations were derived for AEZ IV because of a lack of weather data and for AEZ V because of low‐ yield potential in the DS. These AEZ‐based fertilizer recommendations were embedded into ICM recommendations, with emphasis on weed control and soil fertility management. In addition to optimum fertilizer use, the selection of nutrient‐eYcient rice varieties, adequate radiation and temperature, good water control, and optimum crop management are critical to maximize rice yields and nutrient‐use eYciency (Balasubramanian et al., 2004).

Table XI Forms in Soils, Sources, Critical Values for Deficiency and Toxicity, and Application Methods for Essential Nutrients in Rice Cropping

N

P

K

S

Zn

Fe

Occurrence in soil

Critical value: deficiency

Critical value: toxicity

NHþ 4 , NO3 ; organic matter

1.4‐mg m2 leaf area

–

Solution P $ Labile P $ Nonlabile P; organic P Solution K $ Ex. K $ Non‐ex. K $ Mineral K; organic K SO2 4 , S , elemental S; organic S

Bray‐1 P: 7 mg kg1 soil; Olson‐P: 5 mg kg1 soil Ammonium acetate‐K: 20‐cmol kg1 soil

Excess P induces Zn deficiency in basic soils

Zn in clay minerals; ZnS; Zn adsorbed on amorphous oxides and mg‐carbonates; humic Zn Fe3þ, Fe2þ

Nutrient sourcesb

Methods of applicationc

Urea, DAP, AS, CAN, coated urea; organic/ green manures Rock‐ or single/double/ triple super‐phosphate, DAP KCl, straw or straw compost

Soil two to four splits; LCC based; deep placement; 0.5% urea spray Basal, seedling root‐dipping in soil‐P slurry; OPM‐/ APM‐based rates Basal at low yield level; two splits at high yield level; OPM‐/APM‐based rates

Ca(H2PO4)2 S: 9‐mg >0.07 mg liter1 of H2S kg1 soil (combined (mainly in sandy, peat, and acid sulfate soils) w/Zn deficiency in low OM, high‐allophane materials and oxides, and sandy soils) 1‐N NH4‐acetate (pH Nil 4.8)‐Zn: 0.6‐mg kg1 soil; 0.05‐N HCl‐Zn: 1.0‐mg kg1 soil

Rain or irrigation water; AS, single super‐ phosphate; gypsum, elemental S

Soil application of S‐containing materials

ZnSO4, ZnO

25‐kg ha1 ZnSO4 applied on soil surface/ floodwater; seedling root‐dipping in 2% ZnO solution

2 mg liter1 (dryland rice only)

Fe‐sulfate, Fe‐citrate

Foliar spray of 0.1–0.5% solution; apply fast‐ decomposing GM

–

300‐mg liter1 Fe2þ (aggravated by poor supply of K, P, Zn)

103

(continued)

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Nutrientsa

104

Table XI (continued)

Nutrientsa Mn

Mo

B

Si a

Acid soil: 10‐ to 100‐mg liter1 Mn2þ in soil solution; Calcareous: 0.5‐mg liter1 Mn2þ Cu‐oxides, carbonates, silicates, and sulfides; chelated‐Cu Mo on silicates; Mo‐sesquioxides; organic Mo; MoO2 4 soluble ions H3BO3 or borates; B‐sesquioxide and illite/vermi‐culite clay; organic B Silicates; clay minerals

Critical value: deficiency DTPA þ CaCl2‐Mn: <1 mg liter1

Critical value: toxicity >100‐mg liter1 Mn2þ

– 0.05‐N HCl‐Cu: 0.1‐mg kg1 soil (mostly in peat soils) – 0.15‐mg liter1 soil solution (rarely deficient)

Nutrient sourcesb

Methods of applicationc

Mn‐oxide (flooded rice); Mn‐sulfate (dryland rice)

Soil; foliar spray; soaking of seeds in Mn solution

Cu2O, CuSO4

Ammonium molybdate

Soil or spray application of Cu2O or CuSO4 once in 3 years Rarely applied for rice

Hot water‐B: 0.5‐mg kg1 soil (rarely deficient)

>5‐mg kg1 soil (only in saline, alkaline soils)

B‐rich groundwater (>2 mg B liter1); borates

Rarely applied for rice (applied for wheat in rotation with rice)

mainly in acid sandy soils

–

Burnt rice hull; silicates

Soil application

N, nitrogen; P, phosphorus; K, potassium; S, sulfur; Zn, zinc; Fe, iron; Mn, manganese; Cu, copper; Mo, molybdenum; B, boron; Si, silica. AS, ammonium sulfate; CAN, calcium ammonium nitrate; DAP, diammonium phosphate; KCl, potassium chloride (muriate of potash); rock‐P, rock‐ phosphate. c APM, addition plot method; LCC, leaf color chart; OPM, omission plot method. Information sources: Balasubramanian et al. (1999), Diamond (1985), Dobermann and Fairhurst (2000), Dobermann et al. (2004), Fairhurst and Witt (2002), and Neue and Mamaril (1985). b

V. BALASUBRAMANIAN ET AL.

Cu

Occurrence in soil

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Table XII Indigenous N Supply, Applied Fertilizer Rates, Grain Yield, and N Recovery EYciency of Irrigated Rice in Farmers’ Fields in the Senegal River Delta of West Africa, Based on Field Surveys 1995–1997 Sites

Parameter Season (WS)a No. of cases Mean INSd Applied N (kg ha1) Applied P (kg ha1) Applied K (kg ha1) Grain yield (Mg ha1) RENe (%)

Tiagar 1, Senegal

Tiagar 2, Senegal

Guede´, Senegal

Location, Mauritania

1995b 10 72 101 22 0 4.9 30

1995c 10 60 80 15 7 4.1 38

1996 20 31 117 21 0 5.6 44

1997 37 32 115 20 0 4.4 33

a

WS, wet season. Single‐rice cropping. c Double‐rice cropping. d INS, indigenous N supply. e REN, recovery eYciency of applied N (% of applied N recovered in grain þ straw). Adapted from Haefele and Wopereis (2004). b

For irrigated wetlands with adverse soil conditions such as salinity, alkalinity, acidity, or Fe and Al toxicity, an integrated approach is needed—by combining resistant or tolerant rice varieties with balanced plant nutrition and crop management. For example, ridge planting of rice and the application of balanced (N, P, K, Zn) fertilizers can reduce Fe toxicity significantly in the poorly drained wetlands of rainforests and Guinea savanna zones of West and Central Africa (Sahrawat, 2004a). In semiarid areas, the salinity problem can be tackled by planting salt‐tolerant rice varieties and applying a full dose of recommended fertilizers (WARDA, 2002). 2.

Toposequence‐Based Nutrient Management in an Inland Valley

Soil fertility problems vary according to position along the inland valley toposequence. For coarse infertile soils in the upper and middle parts, fertilizers must be combined with soil amendments such as composts, manure, or rock‐P to optimize rice yields and nutrient‐use eYciency. On the other hand, soils in the valley bottom are relatively high in clay and OM contents and water‐holding capacity, but poor in P and N; therefore, they generally respond to P and N fertilizers, with less additional response to soil amendments. For example, in an inland valley of the Gambia, Jobe (2003) noted that

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rice responded mainly to fertilizers but not to soil amendments such as manure, rock‐P, and phosphogypsum in the valley bottom, whereas rice in the upper and middle parts benefited most from the application of 50% of recommended nutrients through fertilizers and the rest through 4 Mg ha1 of animal manure. Thus, the combined use of organic and inorganic sources of nutrients is recommended by the ISFM strategy to maximize cereal yields, especially in nutrient‐poor soils of the upper reaches (Vanlauwe et al., 2002).

3. Site‐Specific Nutrient Management In Asia, a strategy called SSNM has been developed for precision nutrient management in irrigated rice (Buresh et al., 2003; Dobermann et al., 2004). SSNM provides an approach for ‘‘feeding’’ rice crops with nutrients as and when needed. Farmers dynamically adjust the application and management of nutrients to crop needs according to location and season. SSNM advocates The optimal use of existing indigenous nutrient sources, including crop

residues and manures Timely fertilizer application to meet the deficit between rice demand for

nutrients and the supply of nutrients from soil and organic inputs (Fig. 8). In SSNM, the leaf color chart (LCC) is used to apply N fertilizer as per the crop’s need during the growing season (real time N management) (Balasubramanian, 2004; Shukla et al., 2004). Rice leaf color is monitored in the field with an LCC from 15 days after transplanting to the booting stage at 7‐ to 10‐day intervals, and N fertilizer is applied whenever the leaf color falls below the chosen critical value.

Figure 8 SSNM approach: ‘‘apply fertilizers as and when needed to fill the deficit between crop need and indigenous nutrient supply’’ (source: Buresh et al., 2003).

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An omission plot technique is used for crop need‐based P and K application. A P‐omission plot—a plot with no added P and the full rate of other nutrients—visually demonstrates to farmers the deficit of P. Similarly, a K‐omission plot—a plot with no added K and the full rate of other nutrients—demonstrates the deficit of K (Buresh et al., 2003; Dobermann et al., 2004). The diVerence in grain yield between a nutrient‐omission plot and a full NPK plot is used to make P and K recommendations, using the formula: ‘‘To produce one metric ton of paddy over the yield of the nutrient omission plot, farmers should apply 15‐ to 20‐kg P2O5 and 30‐kg K2O per ha’’ (Balasubramanian et al., 2003). Other nutrients such as S and Zn are applied as per local recommendations at sites deficient in such elements.

E. WATER MANAGEMENT 1.

FOR

RAINFED

AND

IRRIGATED AREAS

Managing Water for Rainfed Rice Farming

Water is the single most critical constraint in rainfed rice farming. For all rice varieties, the water requirement is the highest during the reproductive phase, that is, from PI to heading; therefore, during this period, rice fields must have enough soil moisture in rainfed drylands or standing water (about 0.05‐m depth) in rainfed or irrigated wetlands. Any moisture deficit (due to drought) at this critical period will seriously aVect spikelet formation and grain filling, with drastic reductions in yield or complete crop failure. For drylands and rainfed wetlands, management options are needed to tackle extreme drought and flood events. First, the duration of rice varieties must be matched with the duration of the wet period of the growing season. Second, the planting date is adjusted in such a way that rainfall is adequate during the period of maximum water requirement (PI to heading). Third, minimum tillage and mulching can be practiced in drylands to conserve moisture at critical periods or to overcome unexpected dry spells (Hatibu et al., 2000; WARDA, 2002). RWH is defined as ‘‘concentrating, collecting, and storing rainwater for diVerent uses at a later time in the same area where the rain falls, or in another area during the same or later time’’ (Hatibu, 2000). There are two types of RWH: (1) in situ RWH where rain is captured where it falls and (2) micro‐ and macrocatchment RWH where the runoV from the catchment area is collected and stored for use in the same or downstream areas during the same or later time. Building farm ponds, small reservoirs, or earth dams across water courses is necessary to store peak floods (Hatibu et al., 2000). RWH and the development of microcatchment storage structures must be given top priority in drought‐ and flood‐prone rainfed wetlands. Water from

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farm ponds can be used to provide 1–2 life‐saving irrigations during periods of drought at critical stages to increase and stabilize rice yields. Once supplementary irrigation is available, farmers will tend to apply additional inputs to enhance rice productivity with reduced risks. In addition, farm ponds will help recharge groundwater in the vicinity, which in turn will revive rainfall‐dependent streams and increase water availability in rural areas. If enough pond water is available in the DS, farmers can diversify into growing small plots of vegetables or other cash crops to diversify their income sources and improve their diets. Further diversification into poultry and livestock is possible if enough water is available to take care of them. Farm ponds can also be used to grow fish. In flood‐prone areas, using reservoirs to store and regulate flash floods can minimize crop damage due to sudden floods (Hatibu et al., 2000). In addition, the provision of drainage is critical to reduce Fe toxicity and stabilize rice production. It is diYcult to drain fields during periods of heavy rains. At other times, farmers are reluctant to drain their rice fields for fear of losing applied N, worries about the arrival of the next rain to reflood their fields, and increased weed infestation under nonflooded conditions. In saline and mangrove swamps, ridge planting is recommended to improve drainage and reduce the adverse eVect of excess Fe and Al and salt (WARDA, 2002). Developing stress‐tolerant rice varieties will help increase and stabilize rice yields in poorly drained wetlands with problem soils. IRRI has developed a breeding line (IR73678‐6‐9‐B) derived form O. sativa cv IR64 O. rufipogon that can tolerate acid sulfate conditions in coastal wetlands; Vietnam released it as national variety AS‐996 in 2002. It can be evaluated in acid sulfate soils of Africa. Genes have been tagged for submergence tolerance (Sub1A) (Xu et al., 2006) and other abiotic stresses (Table IX) for incorporation into popular varieties (Gregorio et al., 2006).

2.

Water Use and Water Productivity in Irrigated Rice Systems

Irrigation is the best means of increasing rice production with modern varieties. Improved water conveyance at the systems level and judicious water use at the farm level are critical to derive maximum benefits from irrigation schemes. Large irrigation projects, especially those in the public sector, have often failed in Africa because of several factors: the lack of a well developed and coherent irrigation subsector policy and strategy; high capital and operating costs (especially for irrigation schemes based on pumping); poor cost‐recovery and lack of funds for management; poor operation, repair, and maintenance

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of irrigation infrastructure; inadequate farm support services; poor ownership by users; and low economic viability (FAO‐Aquastat, 2005). Owing to inadequate drainage, salinity and alkalinity have developed in many irrigated areas in the semiarid savanna, Sudan, and Sahel zones of West Africa (Massoud, 1977; WARDA, 1999). Large investments needed to rehabilitate such irrigation schemes and to improve systems level water delivery may not be available from governments. This is where strong public–private partnerships need to be developed to encourage initial investments in reconstructing irrigation infrastructure and providing processing and marketing facilities for the commercialization of irrigated rice production. Farmer‐controlled tube‐well irrigation systems are highly successful in Asia (e.g., in Bangladesh) (Hossain et al., 2006). Such small irrigation schemes have been developed for rice cultivation in the high‐altitude marshlands of Rwanda and Burundi. It is predicted that Africa will have more small‐scale irrigation schemes in the next 25 years. The development of small irrigation schemes, the construction of related rural infrastructure, and the formation of viable water‐user associations must go hand‐in‐hand to realize the full potential of these irrigation projects and to improve farmers’ livelihood. In addition, both land and water managements must be integrated to maintain productivity at a high level and environmental quality at an acceptable level. We need to consider cropping systems instead of individual crops for enhancing water productivity. Barker et al. (1998) show how diVerent practices reduce various components of water losses in irrigated rice systems (Table XIII). Salinity management is critical for improving the productivity of irrigated rice in semiarid areas of West Africa. Among the management practices that minimize salinity problems, crop selection and rotations, land leveling, and salt leaching have long‐term eVects, while tillage, fallowing and mulching, landform and planting method, and the application of manures, fertilizers, and amendments have only a short‐term eVect (Massoud, 1977). Crops such as rice that require frequent irrigation can reduce salinity eVectively if adequate drainage is available. With land leveling and tillage, precautions must be taken not to bring a salt layer, if present, to the crop root zone; any tillage that improves internal drainage will help reduce salt accumulation. Evaporation from exposed soil surface during fallow periods will lead to salt accumulation, and, in this case, mulching can reduce evaporation and thus salt accumulation on the surface. Although the application of organic residues and manures improves soil physical and chemical properties, they carry some salts with them. Proper fertilizer use will not aVect salinity. Amendments are added mainly to sodic soils to reduce the exchangeable sodium percentage and improve permeability and drainage (Massoud, 1977). In addition, eVective management of waste water and saline water for irrigation is also important.

110

V. BALASUBRAMANIAN ET AL. Table XIII Practices for Increasing Water Productivity in Irrigated Rice Farming

Practice

Ta

Improved short‐duration rice varieties Improved agronomic management Changing schedules to reduce evaporation Reducing water for land preparation Changing rice planting practices Reducing crop growth water Making more eVective use of rainfall Water distribution strategies Water recycling and conjunctive use

þ þ

Eb

þ þ þ þ þ

S&Pc

SROd

þ þ þ þ þ

þ þ þ þ þ

RCLe

þ

a

T, transpiration. E, evaporation. c S&P, seepage and percolation. d SRO, surface runoV. e RCL, recycled water use. Adapted from Barker et al. (1998). b

Among the water‐saving technologies, intermittent irrigation or alternative wetting and drying (AWD) is the most promising because it oVers high water productivity coupled with a low penalty on grain yield. In China and the Philippines, AWD is reported to save from 13 to 30% of irrigation water at the field level, with no significant reduction in yield (Cabangon et al., 2001). The development of rice varieties with short duration, early seedling vigor, and tolerance of submergence will increase water‐use eYciency. Several medium‐duration (120–130 days) high‐yielding rice varieties are available for irrigated wetlands: BG 90‐2 and 380‐2; BKN 7033 and 7167; BR 51‐46‐5; IR8, 22, 46, 54, 58, and 841; ITA 123; Sahel 108, 201, and 202; TOX 728‐1; and WITA 1, 3, and 7. However, improved very early maturing rice varieties are needed for irrigated lands in northern Guinea, Sudan savanna, and Sahel zones and early maturing and Fe toxicity‐tolerant varieties for irrigated rice in rainforest zones of SSA (WARDA, 2002). In Asia, hybrid rice varieties were found to be more adapted to water‐deficit conditions because of their early seedling vigor and vigorous root system that favor the eYcient use of available water (Virmani, 1996). But suitable hybrid rice varieties have yet to be developed for SSA. Salinity‐tolerant high‐yielding varieties such as ITA 212 (FARO 35) and ITA 222 (FARO 36) are available for irrigated areas with saline soils. Aerobic rice is another alternative for water‐deficit areas (Atlin, 2005); in this system, aerobic rice varieties grow in nonflooded soils and yield as high

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as 5–6 Mg ha1, using 40–50% less water than flooded rice systems. Water productivity was as high as 0.6–0.8 g grain kg1 water for aerobic rice varieties in North China (Yang et al., 2005). Promising aerobic rice varieties are Apo and Magat for the Philippines (Bouman et al., 2005) and HD 502 and HD 297 for China (Bouman et al., 2006). Research is ongoing for further improvement of rice varieties, crop and water management, and sustainability for aerobic rice systems. In SSA, aerobic rice varieties can be tried in water‐deficit irrigated areas or rainfed wetlands where supplementary irrigation is available.

F. WEEDS, INSECT PESTS, 1.

AND

DISEASES

AND

THEIR MANAGEMENT

Weeds, Their Management, and Their Multiple Uses

Weeds are plants out of place. They are the most serious biological constraint to rice production in SSA (Johnson, 1997). The major rice‐field weeds that occur in SSA are listed in Table XIV. Weed infestation is serious wherever water control is poor as in most rainfed and poorly managed irrigated areas. Improper land preparation and poor leveling lead to poor spread of water and more weeds in irrigated rice fields. Becker and Johnson (1999) observed that Echinochloa species are the most common weeds in fully irrigated systems and Panicum laxum in poorly irrigated fields in West Africa. More weeds germinate and compete with rice under direct seeding than in transplanting conditions (Balasubramanian and Hill, 2002). When soil is depleted with reduced fallow periods, weeds become more competitive than rice, especially in drylands. The parasitic weed, Striga spp. (S. aspera and S. hermonthica), is a major problem in dryland rice mainly in the savanna zone but not in the forest zone (Adagba et al., 2002). Rice yield loss due to weeds can be as high as 25–40%, with total crop loss in extreme cases (WARDA, 1998). Manual weeding is the most common, as herbicides are beyond the reach of poor farmers and, hence, basically suitable for the low‐income countries of SSA. In direct‐seeded irrigated rice trials in the Senegal River delta, weeds had to be controlled until 32 days after seeding (DAS) in the WS and 83 DAS in the DS to get 95% of the rice yield of weed‐free fields (Johnson et al., 2004). Manual weeding, however, is time‐consuming and labor‐intensive. An estimated 27–37% of the total labor input goes for weeding alone (WARDA, 1998). However, timely hand weeding not only controlled weeds eVectively but also maintained higher levels of pest predators such as spiders and beetles in rice fields than in herbicide‐treated plots. Placing the uprooted weeds in piles within rice fields resulted in larger populations of spiders than any other method of weed disposal, including removal from fields (Afun et al., 2000).

112

V. BALASUBRAMANIAN ET AL. Table XIV Major Rice‐Field Weeds Reported from Africaa

Grasses

Sedges

Broad‐leaf types

Parasitic (dryland rice)

Cynodon dactylon Cyperus diVormis Alternanthera sessilis (L.) Striga asiatica (L.) Pers., L., C. happen L., L. Kuntze, Striga R. Br. ex DC., Dactyloctenium C. iria L., aspera (Willd.) Aeschynomene indica aegyptium (L.) Willd., Fimbristylis Benth. L., Ageratum Digitaria spp., ferruginea (L.) conyzoides L., Echinochloa crus‐galli Vahl., F. pilosa Amaranthus spinosus L., (L.) P. Beauv., E. (L.) Vahl., Pycreus Ammania baccifera L., colona (L.) Link., E. macrostachyos, Commelina benghalensis stagnina (Retz.) P. Kyllinga L., C. diVusa Burm. f., Beauv., E. pyramidalis squamulata Euphorbia heterophylla L., Eleusine indica (L.) (Thon), L., Diplazium sammatii Gaertn., Imperata Abildgaardia (Kuhn), Eclypta cylindrica (L.) hispidula prostrata (L.) L., Ludwigia octovalvis Raeuschel, Ischaemum (Jacq.), Trianthema rugosum Salisb., portulacastrum L., Leersia hexandra Sw., Portulaca oleracea L., Oryza sativa (red rice) Zaleya pentandra L. (L.), Panicum repens L., Panicum laxum Sw., Paspalum scrobiculatum L., P. polystachum R. Br., Rottboellia cochinchinensis (Lour.) Clayton, R. exaltata, Setaria pumila (L.) P. Beauv., Spermacoce ruelliae a

Data sources: Adagba et al. (2002), Johnson (1997), and Rao et al. (2007).

Developing weed‐competitive rice varieties is one strategy for cost‐ eYcient weed control in rice crops. In West Africa, O. glaberrima was found to be more tolerant of Striga (WARDA, 1998) and other weeds (Johnson et al., 1998) than O. sativa. As discussed earlier, the dryland NERICA (a cross between O. glaberrima and O. sativa) varieties are highly competitive with weeds due to their early seedling vigor, fast canopy development, and droopy lower leaves that shade out weeds (Futakuchi and Jones, 2005; Haefele et al., 2004; WARDA, 2003–2004). The newly developed wetland NERICA varieties have weed‐suppressing ability similar to that of their dryland counterparts (WARDA, 2003–2004, 2004–2005). The second strategy is planted legume fallows, which are reported to smother

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113

weeds and fix N for the following rice crops. The rice–legume fallow rotation was more productive than the rice–natural fallow rotation (Akanvou et al., 2001a; Carsky et al., 2001; WARDA, 1998). The third strategy of rice seed priming for fast germination and early canopy development in dry direct‐ seeded rice and their eVect on weed suppression is yet to be assessed in the field (WARDA, 1998). Although weeds compete with rice crops, some of them are useful to subsistence farmers. For example, weeds such as Echinochloa colona and E. pyramidalis have multiple uses as fodder, thatching material, and grains; Paspalum scrobiculatum as fodder, poisonous raw grains, and medicine (Burkhill, 1994); Setaria pumila as fodder and thatching material; Trianthema portulacastrum as food and medicine; and Zaleya pentandra as fodder for horses and donkeys (Burkhill, 1985). Other rice‐field weeds such as Rottboellia exaltata and Spermacoce ruelliae provide fodder or grazing matter for livestock.

2.

Major Insect Pests and Diseases and Their Management

Pink stem borer (Sesamia calamistis) is the main pest of dryland rice, whereas AfRGM (Orsylia oryzivora) is common in rainfed and irrigated wetlands. Other pests are striped (Chilo zacconius), white (Maliarpha spp.), and yellow (Scirpophaga spp.) borers and stalk‐eyed fly (Diopsis macrophthalma); grain‐sucking bugs (Aspavia sp., Stenocoris claviformis); case worm (Nymphula depunctalis); and whorl maggot (Hydrellia sp.) (John et al., 1986; WARDA, 2002). Incidence of stem borers and stalk‐eyed fly is severe in all humid and dry zones, while that of grain‐sucking bugs, case worm, whorl maggot, and AfRGM is more severe in humid forest and Guinea savanna zones than in the Sudan savanna zone (John et al., 1986). The major diseases of rice in Africa are blast (caused by Pyricularia oryzae), glume discoloration (fungal complex: caused by Sarocladium sp. and Curuvularia sp.), RYMV, sheath rot (caused by Sarocladium sp.), leaf scald (caused by Rhyncosporium oryzae), sheath blight (caused by Thanetophorus cucumeris), and bacterial leaf blight (caused by Xanthomonas campestris pv. oryzae) (John et al., 1986; WARDA, 2002). Other pests such as rodents and birds attack rice in all ecosystems. Nematodes and termites are a serious problem in dryland rice in some areas. Integrated pest management (IPM) is a decision‐support system for the selection and use of pest control strategies that minimize dependence on chemical pesticides and improve human health and environmental quality. Growing a healthy crop is the key to good IPM. Other IPM technologies for rice are: (1) deployment of pest‐resistant varieties, (2) no early spraying against leaf folders and thrips, (3) an active barrier system for rat control,

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(4) silica application for blast control, and (5) timely and judicious use of fast‐acting bio or synthetic pesticides when pest infestation is serious, threatening the crops. Developing resistant varieties has been the main focus of research at WARDA. Two rice varieties, WITA 8 and WITA 9 (WARDA, 2002), and three wetland NERICAs have been selected for resistance to RYMV (WARDA, 2003–2004, 2004–2005). Breeders employ biotechnology tools to incorporate specific genes into high‐yielding rice varieties to make them resistant to or tolerant of various pests and diseases. For example, two brown planthopper (BPH ) genes for BPH, six Gm genes for gall midge, eight Xa genes for bacterial blight, and eight Pi genes for blast have been tagged at IRRI for incorporation into high‐yielding rice varieties through MAS. Similar work is ongoing at WARDA to identify and tag resistance genes from cultivated and wild rice species to deploy against Africa‐specific insect pests and diseases.

G. GRAIN QUALITY MANAGEMENT: FROM BREEDING

TO

MILLING

As early as 1985, Buddenhagen (1986) pointed out the importance of having good grain quality for local rice in Africa. He gave four reasons: Rice consumers, especially in urban areas, used to high‐quality and low‐cost

imported rice will demand the same level of quality and cost‐eVectiveness from local rice markets The earlier introduced rice varieties in the Sahel of West Africa were high‐ quality, long‐grain rice varieties from Southeast Asia Although poorly milled and maybe not of long‐grain type, many of the old land races of West and East Africa have very good taste (some with good aroma) and cooking quality preferred by local consumers African rice traders are as sophisticated as any in the world to exploit consumer preferences for quality rice. Buddenhagen (1986) recommended the breeding of medium‐grain types with good taste and cooking quality such as Sierra Leone’s Ngovie or LAC 23, or Tanzania’s aromatic rice for rainfed drylands and wetlands, and long‐ grain types for irrigated rice in SSA. Proper plant nutrition and good water and pest control during crop growth are important for producing healthy, well‐filled grains. Rice farmers in Africa lose 15–50% of the market value of grains because of improper handling of rice during and after harvest. To reduce grain losses and maintain grain quality, rice must be harvested immediately after maturity (95% mature), threshed soon after harvest, cleaned, and dried to <14% moisture content (MC). Simple dryers and sealed storage options are

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available for drying and storing the grain properly. Airtight storage is recommended to kill storage pests without pesticides, preserve grain quality, and maintain seed viability. Rice must be milled at 12–14% MC depending on grain types in modern mills to maximize head‐rice recovery. IRRI has developed appropriate and cost‐eVective tools such as a low‐cost moisture meter to monitor grain moisture while processing, super bags for sealed grain storage, a rice‐milling chart to guide optimum milling of rice, and a portable grain quality test kit (IRRI, 2005). An improvement in harvest and postharvest processing will enhance the market value of rice and thus not only improve farmers’ income and livelihood but also enhance profit to millers.

H.

DIVERSIFICATION

OF

RICE FARMING SYSTEMS

Generally, rice is not highly profitable in many areas. However, in the WS, rice is the most suitable crop for wetlands. In the DS, dryland crops such as grain legumes, vegetables, onion, garlic, watermelon, and cotton are grown in Asia and Africa to enhance farmers’ income. In a 5‐year (1996–2001) crop rotation trial in the Accra Plain of Ghana, Nyalemegbe et al. (2003) found that rice–vegetable (rice–okra and rice–eggplant) systems are more profitable than rice–rice, rice–cowpea, or rice–watermelon systems. Vegetable yields were low but the prices were high in the DS and vice versa in the WS. In this study, deep‐rooted crops such as okra and eggplant tapped nutrients from the subsoil, aerated the soil, and added OM to the system, while grain legumes such as cowpea fixed N and added crop residues to improve soil fertility. The rice–rice system reduced sedges considerably, but not the grasses, due to continuous submergence, while the rice–cowpea and rice–okra systems reduced noxious water‐loving grass weeds such as E. crus‐galli and Ischaemum rugosum. Crop‐livestock integration with a cut‐and‐carry fodder system will improve cash flow and generate income for dryland farmers (Otsuka and Yamano, 2005). This system will also improve the use of crop residues as fodder for animals and animal manure to improve soil fertility in rice fields.

I. ICM FOR RICE In the literature, there is no universal definition of ICM. In practical terms, it means good agronomy or good crop management. We use the ICM approach to promote the combined use of adapted rice varieties and location‐specific crop management technologies to increase land and water productivity and profit to farmers in irrigated rice farming (Balasubramanian et al., 2005; Chandrasekaran et al., 2004). From 2004, the FAO of the United

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Nations has adopted the ICM approach to identify and deploy technologies for increasing rice productivity in Asia and elsewhere. ICM has two sets of options: core and location specific. Core options are those that perform similarly in multiple locations, for example a locally adapted variety, good‐quality seed, robust young seedlings, crop need‐ based nutrient application, and IPM. Location‐specific options are those that perform diVerently at diVerent locations or that can be practiced only in certain locations; plant spacing, intermittent irrigation, the use of organic manure, and so on are location specific. Both types of ICM options have to be selected to meet farmers’ needs in each location; they must also be updated periodically to incorporate new research findings as they become available. Additionally, timely harvest and threshing, proper cleaning and drying of grain to 14% MC, the use of sealed storage systems, and proper milling with the help of a rice‐milling chart will reduce grain losses and increase the market value of milled rice. Key components of ICM successfully validated in Asia are: (1) the selection of locally adapted rice varieties, (2) the use of good‐quality seed at low seed rates, (3) the preparation and planting of young seedlings at the four‐leaf stage, (4) transplanting of 1–2 young seedlings per hill in a square pattern (0.2 m 0.2 m to 0.25 m 0.25 m), (5) 2–3 mechanical weedings þ soil stirring at 10‐day intervals from 15 days after transplanting, (6) intermittent irrigation during vegetative and grain‐maturing phases and continuous shallow flooding from PI to flowering, (7) crop need‐based nutrient management (SSNM), and (8) IPM. ICM pilot studies conducted in Asia during 2000–2004 demonstrated that ICM fields produced 1.1–2.5 Mg ha1 more rice yield than rice crops under recommended conventional practices; additional profit due to ICM ranged from US$115 to US$265 ha1 per season (Balasubramanian et al., 2005). Growing healthy crops through ICM will also help reduce pesticide use and pesticide‐related health risks to farmers. ICM will be especially beneficial to smallholders with limited land and suYcient labor. Seed producers will benefit more than grain producers when they adopt the ICM approach. ICM for irrigated rice in West Africa focuses on land preparation, seed rates and sowing time, cultivar choice, and crop establishment techniques, and provides a farming calendar for a given combination of site, sowing date, cultivar choice, and crop establishment technique (WARDA, 2000). Additional ICM options are fertilizer rates for specific target yields, weed and water management techniques, and postharvest techniques (Defoer et al., 2002). In on‐farm evaluations in Senegal and Mauritania, mean rice yield increased from 3.8 Mg ha1 for the farmers’ practice to 5.5 Mg ha1 for ICM plots and the average net benefit rose from US$215 to US$525 ha1, despite the slightly increased cost for ICM plots. Moreover, the adoption of ICM options tends to reduce risks (Haefele et al., 2000). The ICM options

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most attractive to farmers were improved fertilizer and weed management, adapted high‐yielding varieties, and mechanized harvesting and postharvest technologies. Adding proper land leveling; the preparing and planting of young, vigorous seedlings; intermittent irrigation with soil stirring; and improved postharvest management to rice‐ICM recommendations in SSA may further enhance rice yields and profitability to African farmers, as evidenced in Asia (Balasubramanian et al., 2005). The available best management technologies for rice in diVerent ecosystems of SSA are listed in Table XV. These technologies can be deployed as a basket of options to farmers using the ICM approach.

VIII.

RICE INTENSIFICATION ISSUES AND THOUGHTS FOR THE FUTURE

A. RICE INTENSIFICATION IN RELATION HUMAN DISEASES

TO

VECTOR‐BORNE

Farmers consider flooded wetland rice cultivation risky for health reasons— the fear of contracting wetland‐related human diseases such as malaria and bilharzia. The distribution of these two diseases in diVerent countries of SSA is shown in Table V. Malaria caused by mosquitoes (Anopheles gambiae s.l.) is the single biggest killer disease in SSA; about half a million children die of malaria ever year. The severity of malaria incidence depends on the susceptibility of local people and the preventive measures they use. Research findings indicate that the expansion of rice cultivation to a new area or conversion of single‐rice to double‐rice cropping will not further increase the incidence of malaria in the already endemic humid and savanna zones. However, in new irrigation schemes or with the expansion of existing irrigation schemes to new areas in the Sahel, the risk of malaria may increase in the short term as new people and laborers settle the area. As the settlers acquire immunity, the danger of malaria goes down steeply (WARDA, 1996). Of the more than 200 million people infected by bilharzia or schistosomiasis worldwide, about 10% of them are seriously incapacitated—80% of which are in SSA. The disease is caused by worms known as schistosomes— urinogenital schistosomiasis by Schistosoma haematobium and intestinal schistosomiasis by S. mansoni. Aquatic snails and humans provide alternative hosts for the worms. The disease is prevalent where people are in contact with snail‐infested water—slow‐moving rivers and streams and vegetation banks of lakes. Preventive health measures such as the use of boots and the destruction of snail‐infested floating vegetation as well as timely treatment of

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Table XV Available Best Management Technologies for Rice in DiVerent Ecosystems of SSA Irrigated Practice

Transplanted Short‐duration high‐ yielding varieties; salt‐tolerant for saline areas

Seed quality Nursery

Good seed (certified) Good seedbed and thin sowing Young, robust (four‐ leaf stage) 20–30

Seedling quality Seed rate (kg ha1)a Landform Planting

Water/irrigation management b

Weed competitive, early vigor, fast canopy development, nonlodging Good, primed seed – – 60–80

Rainfed wetland

Mangrove swamps

Dryland

Short duration, weed competitive, drought‐ and flood‐ tolerant

Tolerant of Fe and Al toxicity, submergence, salinity

Weed competitive, acid‐tolerant, and P‐eYcient

Good seedbed and thin sowing Young, robust, tall

Good seedbed and thin sowing Robust, healthy

– –

20–30

60–100

Plowing, forming ridges Transplanted on ridges with two to three rows per ridge Good drainage to mitigate tidal flooding; ridges to facilitate drainage

Plowing, leveling, or terracing Broadcast, drilled in rows, or dibbled

Good puddling and leveling Two to three seedlings per hill at 0.2 m 0.2 m spacing

Plowing or puddling and leveling Broadcast or row‐ seeded on puddled or moist soil

20–30 for TPR and 60–80 for DSR Plowing or puddling and leveling Transplanted or broadcast or row‐ seeded on wet soil

Intermittent irrigation up to PI and then after flowering; 0.05‐m water level from PI to flowering

Saturated soil first 10 days; then intermittent up to PI and after flowering; 0.05‐m flood from PI to flowering

Rainwater harvesting and farm pond, reservoir, earth dam; one to two life‐ saving irrigations at critical stages

Terracing on slopes; mulching to conserve moisture; contour live hedges

V. BALASUBRAMANIAN ET AL.

Variety

Direct seeded

Crop need‐based SSNM; adequate P in saline soils; S and Zn in deficient soils

Weed control

Herbicide, manual

Pest controlc Mechanization

Resistant variety, IPM Animal power and two‐wheel tractor for land preparation; simple manual seeders; rotating hoes for weeding; simple threshers and cleaners Drain 2 weeks before harvesting – Harvest soon after maturity (95% mature) Thresh immediately after harvest; do not leave in the field for drying Clean and dry to <14% moisture content (low‐cost moisture meter to manage grain moisture) Store in airtight containers (e.g., painted mud pots, metal drums, super bags) Mill at right moisture content (12–14% MC) (rice‐milling chart; portable grain quality test kit)

Preharvest Harvest Threshing Drying, cleaning Storage Millingd a

TPR, transplanted rice; DSR, direct‐seeded rice. PI, panicle initiation. c IPM, integrated pest management. d MC, moisture content. b

Herbicide, manual; weed‐competitive variety

Ecosystem and toposequence based for wetland rice Weed‐competitive variety; manual weeding

Ecosystem based for mangrove swamps Manual, timely weeding

P before N; rock‐P to legume fallows, then rice Weed‐competitive variety; mulching; manual weeding

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Nutrient management

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infected people with a single dose of an appropriate anthelmintic drug will control the disease eVectively in all areas, including wetland rice areas (WARDA, 1999).

B.

ENVIRONMENTAL ISSUES RELATED TO RICE INTENSIFICATION

IN

SSA

Rice intensification in SSA must take preventive measures to protect the environment and ensure the long‐term sustainability of rice farming. IRRI has developed five environmental indicators to monitor production, biodiversity, pollution, land degradation, and water (IRRI, 2004). Production: For rice production to be sustainable in the long run, it is

important to preserve the natural resources—land, water, and soil fauna and flora—that support rice production. Combining balanced fertilizer use with adequate weed control will enhance rice yields and maintain soil fertility in rainfed and irrigated wetlands of SSA (WARDA, 1998). Conservation agricultural methods such as zero tillage, planted fallows with legumes, and dryland rice‐cover crop rotations are being developed to reduce soil erosion, conserve soil moisture, and improve soil fertility in drylands (Akanvou et al., 2001a; Carsky et al., 2001; Erenstein, 2003; Somado et al., 2003). Biodiversity: Preserving wild rice species and diverse cultivated rice varieties in situ in fields and outside in Gene Banks is critical to guard against pest outbreaks and to supply desirable genes for future needs. Many of the hardy O. glaberrima lines and varieties would be extinct had their seeds not been collected and preserved in Gene Banks, because African farmers had abandoned them to grow their distant Asian cousins (WARDA, 1999, 2001–2002). We need careful planning based on scientific research to seek a balance between the economic return of land conversion and social value of preserving marshlands. It is argued that the conversion of marshlands to wetland paddies will reduce biodiversity and release a huge amount of carbon accumulated under the ground. Thus, only high‐potential mangrove areas must be identified and developed for rice farming and the remaining mangrove forests preserved in their natural state to save the rich mangrove biodiversity and protect local communities against cyclones and tsunami‐induced invasion of coastal lands. Similarly, there is a danger of the harvesting and irreversible drying of peats in inland valleys and basins. In addition, acid sulfate soils must be kept flooded all the time to prevent irreversible drying into a nonusable resource.

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Pollution: Herbicides are eVective in controlling weeds in all systems, but

they have to be applied at the right time and at the right dose to destroy weeds with minimal impact on the environment. The use of IPM for pest control and SSNM and the LCC for precision nutrient management in rice will help reduce pesticide‐ and fertilizer‐related pollution of water sources (Balasubramanian, 2004; Buresh et al., 2003; Dobermann et al., 2004). Methane emissions from flooded rice fields are an environmental issue (IRRI, 2004). An estimated 5–10% of global methane emissions are attributed to flooded rice systems in Asia (Wassmann et al., 2000). Methane mitigation measures developed during the past 20 years include crop residue management, the addition of various soil amendments, and midseason drainage (Wassmann et al., 2000); these measures can be proactively applied in areas of rice intensification in SSA. Land degradation: In irrigated rice areas of the semiarid savanna and Sahel

zones, the danger of soil salinization and alkalization must be addressed through proper drainage and the application of adequate soil amendments and nutrients to minimize or prevent these types of land degradation (Massoud, 1977; WARDA, 1999). Conservation agricultural methods are needed to prevent soil erosion and soil degradation on drylands. Water: Water is the most precious natural resource on Earth, and large amounts of fresh water are diverted to rice cultivation all over the world. It is therefore important to optimize water productivity by using water‐ saving technologies such as AWD, and water‐eYcient short‐duration and aerobic rice varieties (Bouman et al., 2005, 2006) and hybrid varieties (Virmani, 1996). The use of less water does not have to obligate farmers to use more chemical inputs if we develop suitable row‐seeding and line‐ transplanting methods and mechanical weed control options. In rainfed wetlands, RWH is critical to recharge groundwater, revive small streams and rivers, stabilize rice yields, and diversify farming and income sources.

C. PREPARING FOR THE IMPACT OF CLIMATE CHANGE Climate change has become a major concern worldwide. SSA will also face new burdens due to climate change. In SSA, major symptoms of such a change include the disruption of normal climate patterns over large areas, increasing incidences of drought and temperature extremes, heavy flooding and soil erosion/landslides, and increasing levels of salt stress in both inlands and coastal areas. For example, in Mali, the rainy season is now too short for rice and the DS is too hot for potatoes. With time, these problems occur more frequently and with increasing severity, threatening the livelihoods of

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resource‐poor farmers. Therefore, underlying processes and the short‐ and long‐term directions of such changes need to be well understood to formulate appropriate adaptive strategies and policies for farming under the impact of changing climate. Solutions require multidisciplinary and integrated approaches to develop a combination of improved rice germplasm with tolerance of prevailing and anticipated stresses and adapted crop management as well as mitigation and amendment strategies that could help simultaneously protect farmers, consumers, and the environment.

D. TECHNOLOGY DELIVERY AND DEPLOYMENT ISSUES The new knowledge and technologies must be evaluated in producers’ fields and promoted widely for real impact on productivity and livelihood of the poor. The technology deployment part is as important as the supply side of the research‐innovation system. It is important to train, equip, and motivate field and extension staV of the public and private sectors, nongovernmental organizations, and civil society as well as farmer leaders, and to strengthen the research‐extension‐farmer linkage to eVectively move new research findings and technologies from research stations to producers. Training of and technical support to farmers on new technologies, field demonstrations, the organization of farmers’ days at harvest time, and encouragement of farmer‐to‐farmer communication are eVective technology deployment tools. With the advent of information and communication technology (ICT), we can now reach out to farmers scattered far and wide. IRRI has developed a Rice Knowledge Bank (www.knowledgebank.irri.org/)—a virtual digital extension and training information tool for the extension of rice technologies. The scheme for the ICT‐based knowledge delivery system shown in Fig. 9 will help to eVectively deploy rice technologies to producers and collect farmers’ feedback. The development of a region‐specific Rice Knowledge Bank for SSA is important for the eVective deployment of rice technologies. The voluntary organization of farmers into groups is essential to eVectively coordinate farming operations and allocate community resources (water, grazing land), shape supportive farm policy (crop insurance, minimum support price, and so on), develop processing and value addition enterprises, and take up direct marketing. In addition, training, extension, and delivery of knowledge are more eVective with farmer groups than with individual farmers. Private sector‐mediated and supported farmer groups will enable the timely delivery of the latest technologies with required inputs and technical support; guarantee farm credit through commercial banks; develop processing, storage, and marketing facilities in strategic locations; assure buyback of produce from farmers at a predetermined price in each

INCREASING RICE PRODUCTION IN SUB‐SAHARAN AFRICA Rice knowledge bank

Production guides Fact sheets Diagnostics Training resources E-learning Databases Image library

Delivery mode

Rural Cyber units - Internet - CD-based

Pilot sites

Village 1

Village 2 HAM radio SMS phone Mobile ICT vans

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Feedback/ experiences/ derivative products

Village 3

Village n

Figure 9 An example of the ICT‐based knowledge delivery system.

season; and ensure a good‐quality rice supply to national, regional, and global markets.

E. POLICY SUPPORT FOR RICE INTENSIFICATION

IN

SSA

It must be emphasized here that the advent of high‐yielding rice varieties has triggered subsequent changes in supporting policies, such as investments in irrigation, initiation of credit programs, and the establishment or strengthening of national research and extension systems in Asia as the rates of return to such investments increased significantly (Barker and Herdt, 1985; Hayami and Kikuchi, 1982; Otsuka and Kalirajan, 2005, 2006). The national governments in the major rice‐producing countries of SSA must design supportive policies for science‐based intensive rice production systems based on estimated rates of return to various investments. In particular, urgent analysis is needed on the benefits of rehabilitation of irrigation facilities, initiation of farm credit, creation of rural infrastructure and marketing facilities, and development of public–private partnerships. The NARIs in many African countries must be revamped and supported to do relevant research on the emerging issues of intensification of rice farming. Similarly, the public and private sectors, NGOs, and community extension services must be trained, equipped, and enabled to provide focused advice and technical support to rice farmers, small and large. We believe that if

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technological breakthroughs are supported by political will, all these required changes for a Green Revolution in SSA will take root, as occurred earlier in Asia.

IX.

CONCLUSIONS: CHALLENGES TO AND OPPORTUNITIES FOR ENHANCING RICE PRODUCTION IN SSA

The cost of irrigated rice production is high in many SSA countries (FAO‐CORIFA, 2005), mainly because of the high initial investment in irrigation infrastructure and the poor operation of many irrigated rice schemes (FAO‐Aquastat, 2005). In addition, irrigated rice farmers have not realized the full potential of improved irrigated rice varieties due to less‐than‐optimum input use and crop management. On the other hand, rainfed wetland rice production by smallholders is often constrained by many biotic and abiotic stresses as well as inadequate crop management— all resulting in low yields, less than 1.5 Mg ha1. Moreover, postharvest losses are high and the quality of milled rice is often poor in many countries. Therefore, the production of locally preferred rice at a competitive price is the biggest challenge to African farmers. The most important challenge to developing high‐yielding rice varieties with acceptable grain quality and resistance to or tolerance of local pests in rainfed and irrigated wetland ecosystems is being addressed progressively. Although O. glaberrima Steud (African rice) is native to SSA, Asian rice (O. sativa L.) introduced during the early sixteenth century by European and Asian traders is cultivated in many irrigated areas of SSA (Buddenhagen, 1986). Breeders are trying to incorporate tolerance of drought and local biotic stresses into the introduced high‐yielding‐irrigated rice varieties. Of course, combining high yield with grain quality is critical to transforming subsistence agriculture into commercial rice farming. The second challenge lies in identifying, branding, and promoting high‐ quality locally adapted rice varieties in national, regional, and international markets. Although increasing crop yields is very important, a major problem faced by all SSA rice producers is to reduce postharvest losses, which presently account for 15–50% of the market value of production. We believe that to address these challenges, it is critical to improve the weak rice R&D capacity in SSA. In Mozambique, for example, only 2–3 national scientists are available to address the problems of 500,000 rice farmers who grow rice on 200,000 ha of land; in Tanzania, only 23 national scientists are involved in rice research for 322,000 ha of rice land; and, in Kenya, only 5 researchers are employed to examine the problems of more

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than 100,000 rice farmers. Even these limited human and financial resources tend to be spent on irrigated rice schemes. The third challenge, the most critical one, is the absence of a coherent and comprehensive policy, plan, and program to tackle the many constraints and deficiencies of the national rice sector in African countries. It must be emphasized here that the development of high‐yielding rice varieties and profitable production technologies is a prerequisite to trigger changes in supporting policies, such as investments in irrigation, initiation of credit programs, revamping of national rice R&D systems, and development of rural infrastructure and market systems for local rice. Great opportunities exist to increase rice production and strengthen both household and national food security in SSA. First, national governments are trying to increase local rice production to reduce rice imports. Simultaneously, the donor community has doubled its aid to SSA for reducing poverty and improving food security and nutrition. In such an environment, the vastly underexploited rice sector oVers a tremendous opportunity to substantially increase local food production and improve food security and farmers’ income and livelihood. Donors such as the Rockefeller Foundation and the USAID through Borlaug LEAP (Leadership Enhancement in Agriculture Program) Fellowships are willing to support the training of a large number of new African plant breeders and crop production scientists to help develop the agricultural sector on the continent. In addition, the NEPAD’s (New Partnership for African Development) CAADP (Comprehensive African Agricultural Development Program) initiative and various other multilateral facilities can help develop the rice sector in SSA. The WARDA and IRRI have a wide array of international expertise in modern rice breeding, biotechnology, and production and postproduction management. These two centers are joining hands to help improve the rice sector in SSA. This is a great opportunity for African agricultural R&D institutions to improve their rice science and technology‐development capability through focused training and joint development, and the implementation of collaborative rice R&D projects.

ACKNOWLEDGMENTS Earlier review of this chapter and critical comments and suggestions by Dr. K. L. Sahrawat of ICRISAT, India are gratefully acknowledged. The authors thank Dr. Bill Hardy for editorial assistance and improvement of the chapter.

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PHOSPHATE REACTION DYNAMICS IN SOILS AND SOIL COMPONENTS: A MULTISCALE APPROACH Yuji Arai1 and D. L. Sparks2 1

Department of Entomology, Soils and Plant Sciences, Clemson University, Clemson, South Carolina 29634 2 Department of Plant and Soil Sciences, University of Delaware, Newark, Delaware 19717

I. Introduction II. P Chemistry III. Phosphate Adsorption on Soil Components A. Phosphate Adsorption on Soils (Empirical Approaches) B. Phosphate Retention as AVected by Physicochemical Properties of Soils C. pH EVects on Phosphate Adsorption on Variable Charge Minerals D. Phosphate Adsorption on Metal Oxides E. Phosphate Adsorption on Phyllosilicate Minerals F. Temperature EVects on P Adsorption on Soil Components G. I EVects on P Surface Complexation IV. Phosphate Surface Complexation on Soil Components A. Surface Complexation‐Modeling Approaches B. Electrophoretic Mobility Measurement Studies C. Ex Situ Spectroscopic Studies D. In Situ Spectroscopic Studies V. Residence Time EVects on Phosphate Adsorption and Desorption in Soils and Soil Components A. Residence Time EVects Theory B. Slow Adsorption C. Slow Desorption Process and Hysteresis D. Solid‐State, Inter‐, and Intraparticle DiVusion E. Surface Precipitation F. Higher Energy Binding Through Chemical Reconfiguration VI. Future Research Needs References

135 Advances in Agronomy, Volume 94 Copyright 2007, Elsevier Inc. All rights reserved. 0065-2113/07 $35.00 DOI: 10.1016/S0065-2113(06)94003-6

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Macroscopic‐ to more recent molecular scale investigations have enhanced our knowledge of soil phosphorus (P) chemistry, including the retention/ release mechanisms in soils and soil components. Phosphate uptake on metal (oxy)hydroxide and phyllosilicate mineral surfaces and in soils generally increases with decreasing pH. Rapid adsorption kinetics is generally observed on many soil adsorbents at acidic pH. P fixation mechanisms such as inner‐sphere complexation and intra‐ and interparticle diVusion often result in slow P release (i.e., hysteresis and irreversible reactions), creating challenges in remediating agricultural soils with high accumulations of P. This chapter covers some of the historical soil P chemical research findings via macroscopic approaches but focuses on more recent molecular scale approaches for elucidating P retention/release mechanisms. # 2007, Elsevier Inc.

I. INTRODUCTION Phosphorus (P) is an indispensable element for all living protoplasmic organisms because of its genetic role in ribonucleic acid and as an essential nutrient, along with N, for growth. More than 90% of the total P in the soil– plant–animal system is in soils and less than 10% is in the remaining biological systems (Ozanne, 1980). The P content of the lithosphere is 1200 ppm while it is 200–5000 ppm (an average of 600 ppm) in soils (Lindsay, 1979). In the hydrosphere, typical concentrations of total P in domestic wastewater, agricultural drainage, and lake surface waters are 3–15, 0.05–1, and 0.01–0.04 ppm, respectively (Snoeyink and Jenkins, 1980). In the last several decades, excess P has been recognized as point and nonpoint source pollutant throughout the world due to long‐term anthropogenic inputs (i.e., municipal and industrial eZuents, and synthetic and animal‐ based fertilizers) (Parry, 1998; Ryden et al., 1973; Sharpley et al., 2000; Vaithiyanathan and Correll, 1992; Withers, 1996). While point source pollution has been eVectively reduced since the late 1960s, many diYculties still remain in controlling nonpoint P pollution which impacts freshwater and seawater biogeochemical cycles. Eutrophication causes problems in recreational, industrial, and drinking water supplies due to overgrowth of algae and cyanobacteria and their decomposition which leads to a dissolved oxygen shortage (Kotak et al., 1993; Sharpley and Rekolainen, 1997). Consumption of such water, containing cyanobacteria, can be a serious health hazard to livestock and humans due to its neuro‐ and hepatotoxic eVects (Lawton and Codd, 1991). P mobility in agricultural settings, due to surface runoV and leaching, also contributes to a wide range of water‐related problems, including

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summer fish kills. For example, the microscopic organism Pfiesteria (Burkholder and Glasgow, 1997) produces a toxin, killing various lower food chain marine organisms (fish and crustaceans). The toxin may also be harmful to humans because it is known to cause symptoms such as nausea, migraine headaches, skin sores, acute loss of learning, and memory problems in laboratory rats (Levin et al., 1997). Water quality problems as well as P‐related environmental impacts remain despite many attempts to better understand abiotic P chemical processes in soils and sediments. This chapter provides an overview of P research on soils and soil mineral components, which has been conducted at the macroscopic and, more recently, the molecular level scale.

II. P CHEMISTRY P belongs to the Group VA in the periodic table with an electronic configuration of ([Ne] 4s2 4p3). It is stable in the pentavalent state to form an orthophosphate anion (i.e., phosphate) that retains a near tetrahedral complex surrounded by four oxygen atoms. In most soil/water environments, H2 PO 4 and HPO2 4 are the thermodynamically favorable species (pKa1 ¼ 2.1, pKa2 ¼ 7.2, and pKa3 ¼ 12.3). There are several other forms of P‐containing compounds (polyphosphates, metaphosphates, and organic P). The condensed forms of inorganic P (i.e., polyphosphate and metaphosphates) are formed with two or more orthophosphate groups. Whereas polyphosphates are linear O‐P‐O linkages, the metaphosphates are cyclic. The organic P percentage of total soil P can range from 20 to 80% (Dalal, 1977). Several forms of organic P have been identified in soils. They are inositol phosphate, nucleic acids, and phospholipids. Inositol phosphate (phytic acid) makes up more than 50% of the total organic P due to its high stability in soils whereas the phospholipid content comprises as little as 0.5–7% of total organic P (Dalal, 1977). Nucleic acid, which originates from the decomposition of microbes, plants, and animal remains, is the smallest (less than 3%) fraction of the total organic P (Dalal, 1977). P forms complex minerals with a wide variety of elements. About 150 P minerals are known (Cathcart, 1980). According to Povarennykh’s structural classification, P minerals can be placed into four groups. They are framework, insular, chain, and layer minerals. A majority of the P minerals belong to insular minerals, including the apatite group (Ca2Ca3(PO4)3(OH, F)) and wavellite (Al3(PO4)2(OH)3 5H2O). P solubility products are generally controlled by pH, concentration of P, and divalent (e.g., Ca2þ, Mg2þ, and Fe2þ) and trivalent cations (e.g., Al3þ and Fe3þ) in bulk solutions. Where the latter cations are less available, P solubility is strongly controlled by adsorption onto clays and clay minerals in subsurface horizons. In reduced soil environments, P solubility can be increased due to the

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reductive dissolution of iron‐containing adsorbents like iron oxides and ferric‐P minerals (Diaz et al., 1993; Gale et al., 1992; Shapiro, 1957).

III.

PHOSPHATE ADSORPTION ON SOIL COMPONENTS

Surface species (inner sphere or outer sphere) and the P‐bonding environment (e.g., monodentate mononuclear and bidentate binuclear) can be identified and studied using macroscopic, microscopic, and spectroscopic techniques. Electrophoretic mobility (EM) measurements, ionic strength (I ) eVects on the adsorption envelope, and sorption‐proton balance data can be used to indirectly distinguish predominant surface complexes at mineral–water interfaces. X‐ray absorption spectroscopy (XAS) and attenuated total reflectance Fourier transform infrared (ATR‐FTIR) spectroscopy are two powerful techniques that can be employed to directly identify not only the types of surface complexes (inner‐sphere or outer‐sphere) but also the bonding mode at mineral–water interfaces under in situ conditions.

A. PHOSPHATE ADSORPTION

ON

SOILS (EMPIRICAL APPROACHES)

The retention of P in soil–water environments has received much attention in soil chemical studies due to its important biogeochemical role in the environment. It is often found that P adsorption in soils increases with decreasing pH (Sanchez and Uehara, 1980). P adsorption reactions on acidic soils are typically biphasic, characterized by an initial rapid reaction followed by a much slower reaction (Barrow, 1985; Parfitt and Smart, 1978). The slow and continuous P adsorption on soils and soil components has been reported by many researchers (Barrow and Shaw, 1975; Munns and Fox, 1976; van der Zee and van Riemsdijk, 1988; van Riemsdijk and de Haan, 1981). Researchers have attempted to understand P adsorption phenomena using macroscopic approaches (e.g., adsorption isotherms and envelopes) coupled with empirical and surface complexation models. In the past, P adsorption isotherms were extensively fitted to the Langmuir equation to describe diVerent adsorption ‘‘sites.’’ These sites were defined by multiple linear portions of the Langmuir plot (Fried and Shapiro, 1956; Muljadi et al, 1966; Olsen and Watanabe, 1957; Ryden et al., 1977b; Yao and Millero, 1996). Some investigators used a two‐site Langmuir equation to describe the low‐ and high‐energy sites involved in the adsorption reaction. A P adsorption reaction at pH 7 on 19 Vertisols (clay rich soils) was described using a two‐site Langmuir equation. The maximum adsorption

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values from the two‐site model were highly correlated with the calcite content for high‐energy sites and dithionite‐citrate‐bicarbonate (DCB) extractable iron for low‐energy sites (Lo´pez‐Pineiro and Navarro, 1997). Yao and Millero (1996) used a Langmuir adsorption isotherm equation to understand P adsorption on birnessite surfaces (Yao and Millero, 1996). The data exhibited three linear segments, which were related to a variety of adsorption‐binding sites. Veith and Sposito (1977) showed that the Langmuir equation can equally well describe both adsorption and precipitation reactions (Veith and Sposito, 1977). Such empirical equations should not be used to interpret any particular adsorption mechanism or even if adsorption, as opposed to precipitation, actually has occurred (Sposito, 1989).

B. PHOSPHATE RETENTION AS AFFECTED BY PHYSICOCHEMICAL PROPERTIES OF SOILS The retention of P on soils is highly dependent on the physicochemical properties of the soils, for example crystalline and amorphous iron and aluminum oxides, and organic matter (OM), clay and calcium contents. Many researchers have shown a correlation between these parameters and adsorption and chemical extraction data. Long‐term (256 days) P adsorption on 12 Oxisols (well‐weathered soils) was investigated by Barro´n and Torrent (1995). In their experiments, the adsorption maximum from the Langmuir equation suggested that slow adsorption was highly correlated with the ratio of the OM content and the specific surface area of the solids. High P retention in an OM (i.e., an aluminum‐substituted woody peat) was also observed (Bloom, 1981; Lo´pez‐Hernandez and Burnhan, 1974). Several researchers have reported that P retention in various soils (e.g., Entisols, Oxisols, and soil clays) are highly associated with ammonium oxalate extractable iron (Feox) and/or aluminum (Alox), DCB extractable iron and/or aluminum, and Fe oxides detected by Mo¨ssbauer spectroscopy (Arai et al., 2005; Bloom, 1981; Lookman et al., 1995; Lo´pez‐Hernandez and Burnhan, 1974; Maguire et al., 2000; Sei et al., 2002). In the study of Arai et al. (2005), results of sequential inorganic P fractionation indicated that ammonium oxalate extractable P/Al/Fe fractions were predominant in long‐term poultry litter‐amended Southern Delaware agricultural soils, followed by crystalline iron phosphate‐associated fractions, crystalline aluminum phosphate‐associated fractions, soluble P fractions, and calcium phosphate‐associated fractions (Arai et al., 2005). These Al/Fe‐P associations were also supported by results from electron microprobe analyses. High P retention in calcium rich soils and soil components at pH > 7 has also been reported. Kuo and Lotse (1972) observed strong P retention on calcite and Ca‐kaolinite, and a second‐order kinetic equation was fitted to

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obtain the rate coeYcient. The coeYcient was 30,000 times higher for calcite than for Ca‐kaolinite. A similar correlation was also observed for P adsorption on calcite‐rich Vertisols (Lo´pez‐Pineiro and Navarro, 1997). While many macroscopic studies have shown a strong correlation between P retention and specific soil properties, they have not provided any information on the spatial distribution of P in heterogeneous systems. Several microscopic techniques have been applied to investigate P retention phenomena in heterogeneous systems. Phosphate fixation in P fertilizer‐amended soils (pH 7) was investigated using scanning transmission electron microscopy (STEM) along with energy dispersive X‐ray spectroscopy (EDX) (Pierzynski et al., 1990b). In the separated clay fractions, it was observed that discrete P‐rich particles were highly associated with Al and Ca in the highest density‐separated fraction (<2.2 Mg m–3). Furthermore, the EDX model calculations, using elemental ratios (cations/P), also showed an Al and Ca enrichment associated with P. These results were later supported by a P solubility equilibrium study which predicted the formation of varicsite (aluminum phosphate) like solids at pH < 6.8 and calcium phosphate like solids at pH > 6.8 (Pierzynski et al., 1990a). The P mineralogy of Florida soils derived from phosphoric deposits was investigated using X‐ray diVraction (XRD), thermal analysis, and selective dissolution (Wang et al., 1991). The XRD data showed the presence of carbonate‐fluroapatite (Ca10(PO4, CO3)6F2–3), wavellite (Al3(PO4)(OH)3 5H2O), and crandallite (CaAl3(PO4)2(OH)5 H2O). The oxalate extractable P (Pox) was associated with Alox in the surface horizons. Endothermic diVerential scanning calorimetry peaks (90–100 C), which originated from dehydroxylation of amorphous Al‐ and Fe‐P minerals, were eliminated by oxalate extraction. This suggests that P might associate with amorphous Al and Fe minerals. A similar study was conducted on manure‐derived surface soils and sediments (pH 6.9–9.5) (Harris et al., 1994). Amorphous apatite and ferrous‐P minerals (vivianite) were detected by XRD in stream sediment samples. The lack of crystalline Ca‐P minerals suggests that the manure components might inhibit the crystallization of Ca‐P minerals.

C. pH EFFECTS ON PHOSPHATE ADSORPTION ON VARIABLE CHARGE MINERALS Phosphate adsorption on soil minerals is greatly influenced by the pH of the bulk solution. Metal oxides and phyllosilicate minerals in soils contain surface functional groups (unsatisfied bonds with respect to the repeated bonding of the unit cells). Examples include: (1) the siloxane surface associated with the plane of oxygen atoms bound to the silica tetrahedral layer of

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phyllosilicate minerals, and (2) hydroxyl groups that are associated with the edges of inorganic minerals such as phyllosilicates, metal oxides, oxyhydroxides, and hydroxides (Sparks, 1995a). When the mineral is hydrated, the metal ion (Lewis acid) sites are occupied with water molecules and/or are associated with Lewis base sites (hydroxylated surface) because of the dissociative chemisorption of the water molecules. The surface sites can be protonated and deprotonated depending on the pH of the bulk fluid. Therefore, they are often called variable charge (pH‐dependent charge) mineral surfaces. The speciation of oxyanions is also influenced by pH. The dissociation constant (equilibrium constant), Ka, for the oxyanions refers to the reaction in which an acid donates a proton to water. The larger the Ka value, the higher the tendency to donate protons to water molecules. Therefore, oxyanion adsorption on soil components is a function of both the net surface charge density of the adsorbent and the chemical speciation of the adsorbate which in turn are dependent on the pH of the bulk fluid (pHb). In general, P adsorption on inorganic minerals increases with decreasing pHb 2 due to: (1) the negatively charged chemical species (i.e., H2 PO 4 and HPO4 ) and (2) the positively charged mineral surfaces, when pHb < PZC, of the solids.

D. PHOSPHATE ADSORPTION ON METAL OXIDES The first dissociation constant of P is 2.2, this is followed by constants of 7 and 12.8. At most environmental pH values (4–8), the species are predominantly in deprotonated forms (negatively charged species), and the charge properties of metal oxides are positive due to the PZC of the solids [i.e., 6.5–8.5 for iron oxides, 8.2–9.1 for aluminum oxides; an exception is manganese oxides (e.g., birnessite) 2.8]. Therefore, P is expected to adsorb on metal oxide surfaces strongly via electrostatic interaction when pHb‐PZC is less then zero and to predominantly adsorb via ligand exchange when pHb‐PZC is greater than zero. Yao and Millero (1996) studied P adsorption on birnessite surfaces in 0.7‐M NaCl solutions and seawaters. In both media, P adsorption was maximized at pH 3 and gradually decreased with increasing pH (Yao and Millero, 1996). Hingston et al. (1967) reported that P adsorption on goethite increased with decreasing pH. The P adsorption envelope showed inflection points near the pK values of phosphoric acid (Hingston et al., 1967). A similar pH‐dependent adsorption behavior has been observed for P on akaganeite, ferrihydrite, hematite, goethite, boehmite, and amorphous Al(OH)3 (Bleam et al., 1991; Chen et al., 1973; Chitrakar et al., 2006; Shang et al., 1992). The results of our recent investigation agree with the previous research findings. Phosphate adsorption envelope kinetics at the

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Y. ARAI AND D. L. SPARKS 100 PZSE ∼ 8

P sorbed (%)

80

24 h 12 h 4h 2h 1h 30 min 10 min 5 min

60

40

20

0 3.5

4.5

5.5

6.5 pH

7.5

8.5

9.5

Figure 1 Phosphate adsorption kinetics (suspension density, 1.25 g liter1; I, 0.1‐M NaCl; and initial P concentration, 1 mM) at the ferrihydrite–water interface as a function of pH.

ferrihydrite–water interface [five‐point Brunauer‐Emmett‐Teller (BET) surface of two‐line ferrihydrite, 260 m2 g1; suspension density, 1.25 g liter1; I, 0.1‐M NaCl; and initial P concentration, 1 mM] at pH 4.0–9.5 are shown in Fig. 1. Phosphate adsorption increases with decreasing pH from 9.5 to 4 (Fig. 1). Whereas, 98.2% adsorption (2.70 mmol m2) at pH 4 was achieved after 24 h, slow adsorption continued at pH > 4 after 24 h (Fig. 1). Only 58% (1.59 mmol m2) of total P uptake was achieved at pH 9.5 after 24 h and the slow P uptake continued after 24 h (Fig. 1). The surface charge density of ferrihydrite becomes more negatively charged with increasing pH from 4.0 to the PZSE of the solid (pHPZSE 8). The mineral surface at pH < 8 is more positively charged than at pH > 8, and the protonated mineral surfaces at pH 4 strongly attract negatively charged P solution species (e.g., H2 PO 4 ). It is interesting that the extent of P adsorption kinetics rapidly drops after pH 8. This is probably due to changes in net surface charge density of ferrihydrite from positively to negatively charged surfaces that repel HPO2 4 species.

E. PHOSPHATE ADSORPTION

ON

PHYLLOSILICATE MINERALS

While the metal oxides exhibit a strong aYnity for P at acidic pH values, the phyllosilicate minerals show a diVerent adsorption capacity. In general, the PZC of clays is lower than that of iron and aluminum (oxyhdr)oxides (i.e., 4.6 for kaolinite and 2.5 for montmorillonite) (Sparks, 1995a). Therefore, the

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pH‐dependent edge sites of phyllosilicate minerals are generally negatively charged at most environmental pHs (4–8). If there is oxyanion adsorption at the higher pH range, the oxyanions usually specifically adsorb onto edge sites via ligand exchange. Phosphate adsorption on illite and kaolinite gradually increases from pH 3 to 5, and then the adsorption decreases with increasing pH (Chen et al., 1973; Edzwald et al., 1976).

F.

TEMPERATURE EFFECTS ON P ADSORPTION ON SOIL COMPONENTS

Temperature may have two distinct eVects on a chemical reaction. These are: (1) the rate of reaction and (2) the equilibrium end point (Barrow, 1987). Reaction at high temperature results in an increase in the reaction rate and a decrease in the subsequent desorption if the reaction involves activated complexes (intermediate‐ or high‐energy states) (Barrow, 1979b). The eVect of temperature on the chemical reaction can be explained using transition state theory (Eyring reaction rate theory). The schematic reaction flow is: A þ B $ ABz ! Products

ð1Þ

where A and B are the reactants and ABz is the activated complex. Using transition state theory, the elementary reaction process can be expressed as: (1) the total molecular partition functions per unit volume (qi) for the reactant species and for the activated complexes (qz) and (2) the diVerence in zero‐point potential energies between the activated complex and reactants (E0): k¼

ðkB TÞ h

qz E0 exp qA qB ðkTÞ

ð2Þ

where k is the elementary rate constant, kB is the Boltzmann’s constant, T is the absolute temperature, and h is the Planck’s constant (Stumm and Morgan, 1995b). The previous equation can also be thermodynamically reexpressed as

kB T z gA gB K k¼ gz h

ð3Þ

h i z where Kz is the thermodynamic formation constant exp DG and g is RT the activity coeYcient. Since the activation complexes are diYcult to measure using the above equations during temperature‐dependent reactions, the linearized Arrhenius empirical rate law is commonly used.

144

Ea ln k ¼ ln A þ RT

Y. ARAI AND D. L. SPARKS

ð4Þ

where A is the the pre‐exponential factor, Ea is the activation energy, R is the ideal gas constant, and T is the temperature. Combining the transition state theory and the above thermodynamic expression gives A ¼ [(kBT)/h]exp(DSz /R) and Ea ¼ DHz þ RT, assuming the activity coeYcients are all unity. If an endothermic adsorption reaction (positive enthalpy) is observed, one can estimate the apparent activation energy of the reaction. A linear plot between ln k versus 1/T allows one to obtain the value of the slope (Ea/R). The reaction is generally diVusion limited when Ea < 42 kJ mol1, whereas a chemically controlled reaction is suggested when Ea > 42 kJ mol1 (Sparks, 1995b). Various results have been reported on temperature‐dependent P adsorption and desorption on soils and soil components. In acid soils, P adsorption increases with increasing temperature and desorption is subsequently reduced, suggesting that the reaction involves activated complexes. (Barrow, 1979b; Barrow and Shaw, 1977; Chien et al., 1982; Sheppard and Racz, 1984). van Riemsdijk and Lyklema (1980) reported that P adsorption on gibbsite at pH 5 increased with increasing temperature (2, 12, 22, and 45 C). An estimated activation energy of 63 4 kJ mol1 was determined, suggesting that the reaction was controlled more by chemical than by physical processes (van Riemsdijk and Lyklema, 1980). Yao and Millero (1996) reported that P adsorption kinetics on birnessite surfaces were temperature dependent at pH 8. P uptake increased with increasing temperature from 5 to 35 C during the 50 h of the kinetic experiments. Assuming a pseudo first‐ order reaction during the initial 5 min of reaction, they reported an apparent activation energy of 9 kJ mol1. With goethite, however, P adsorption increased with decreasing temperature from 68 to 25 C. This indicated an exothermic temperature‐dependent reaction and nonactivated complex formation (Madrid and Posner, 1979). Temperature‐dependent P desorption kinetics from desert soils were studied using an anion‐exchange resin (Evans and Jurinak, 1976). The energy of activation was estimated from the data based on 4 h of desorption. An activation energy of less than 42 kJ mol1 in all soils was determined, suggesting that P release might depend on diVusion processes. Conversely, Barrow (1979b) found that the activation energy for the P adsorption (forward) reaction was similar to that of the desorption (backward) reaction. The values for both steps were 80 kJ mol1, suggesting that the rate‐limiting steps for both reactions were not diVusion‐controlled. Bar‐Yosef and Kafkafi (1978) estimated activation energies for initial rapid and slow P desorption processes for kaolinite of 67.8 and 20.8 kJ mol1, respectively. This suggested that chemically controlled desorption was followed by diVusion‐controlled desorption.

PHOSPHATE REACTION DYNAMICS

G.

145

I EFFECTS ON P SURFACE COMPLEXATION

I can have an influence on both the rate of the elementary reaction and the type of surface complexation (inner‐sphere and/or outer‐sphere complexation) (Hayes et al., 1988; Stumm and Morgan, 1995b). Oxyanions adsorb onto variable charge mineral surfaces by both inner‐sphere (via ligand exchange) and outer‐sphere complexation (via electrostatic interaction) (McBride, 1989). These surface species can be indirectly inferred by studying the I eVects on the degree of adsorption (Hayes et al., 1988). Inner‐sphere complexes (i.e., selenite) are insensitive to changes in I due to ligand exchange adsorption mechanisms, while outer‐sphere complexes (e.g., selenate) are sensitive to changes in I because of competition from the counteranions of the indiVerent electrolytes. Many macroscopic studies have investigated I‐dependent oxyanion adsorption behavior on clays and clay minerals. Ryden et al. (1977b) reported that P adsorption isotherm on New Zealand loamy soils was not significantly aVected by I eVects ([NaCl] ¼ 0.001–1 M) at <6‐mM equilibrium concentrations (Ceq). Arai and Sparks (2001) showed no I‐dependent P adsorption on amorphous iron oxyhydroxide at pH 4–7.5. Similarly, P adsorption on birnessite at pH 2–8.5 was not aVected by changes in salinity (5–35 ppt) (Yao and Millero, 1996). On the basis of the theory above, these results might indicate the formation of inner‐sphere surface species on the adsorbents. However, there is an exception to this theory. Phosphate adsorption often increases with increasing I. Several researchers have documented the reaction on soils (New Zealand loamy soils) and soil components (kaolinite, montmorillonite, illite, and goethite) (Barrow, 1980; Edzwald et al., 1976; Helyar et al., 1976a,b; Ryden et al., 1977b). Two theories have been suggested to explain this unique adsorption behavior. They are: (1) the quadruple layer model theory and (2) the diVuse double layer theory. Barrow’s data were successfully fitted using a quadruple layer adsorption model (Bowden et al., 1980). Such complex model applications are generally not subject to direct experimental confirmation because they employ several fitting parameters that cannot be analytically measured (McBride, 1997). Decreased double layer thickness (DLT), due to increased I, allows the oxyanions to closely approach the negatively charged surface, and then they adsorb via ligand exchange. Inner‐sphere adsorption, however, should not be influenced by either the DLT or the repulsion forces because the specific adsorption is in direct coordination with discrete surface metal cations (McBride, 1997). Therefore, the quadruple model and DLT are inappropriate to describe the above P adsorption behavior. McBride used the simple mass action principle to explain adsorption phenomena. This approach is similar to the constant capacitance model (CCM) described by Goldberg and Sposito (1984b). It ignores, however, the

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surface electrical potential correction term, c, by assuming that any inner‐ sphere and outer‐sphere counterions are adsorbed to balance the surface charge created in the process. In the case of high I, the negatively charged surfaces created by inner‐sphere P adsorption are likely to be neutralized by coadsorption of cations (i.e., Naþ) from indiVerent electrolytes maintaining the charge balance. This mass action principle favors such reactions when a higher concentration of indiVerent electrolyte is present. No spectroscopic evidence, however, is available to support this theory.

IV. PHOSPHATE SURFACE COMPLEXATION ON SOIL COMPONENTS While numerous macroscopic studies have investigated adsorption behavior with isotherms and envelopes, they have not provided any information on adsorption mechanisms (i.e., surface complexation) at the molecular scale. Recent empirical and modeling approaches [proton balance measurements and surface complex models (SCM)], and microscopic [electrophoretic mobility (EM) measurements] and spectroscopic (FTIR spectroscopy and XAS) studies have provided better insight on oxyanion adsorption mechanisms at the molecular scale.

A. SURFACE COMPLEXATION‐MODELING APPROACHES The lack of mechanistic information obtained from previous empirical approaches (e.g., Langmuir and Freundlich adsorption isotherm equations) has led to the use of SCM to describe adsorption phenomena. SCM are chemical models that are based on the molecular description of the electric double layer using equilibrium‐derived data (Goldberg, 1992). The CCM is one of the many surface complexation models and has been successfully used to describe an array of chemical reactions on mineral surfaces. These include adsorption, desorption, and dissolution reactions on soil components. For instance, the CCM has been used to quantitatively describe P adsorption on 44 noncalcareous soils over a wide range of pH (4.9–7.6) (Goldberg and Sposito, 1984a). Using the derived parameters (intrinsic surface protonation dissociation constants, capacitance density, P packing area parameters), the CCM described the inner‐sphere adsorption mechanisms (i.e., protonated and nonprotonated P surface species) quite well using adsorption envelope and isotherm data. Goldberg and Sposito (1984b) used a similar CCM approach, which consists of adjustable surface protonation–dissociation constants, surface complexation constants, and capacitance density parameters, to predict P adsorption on

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aluminum oxide and iron oxide surfaces. Assuming inner‐sphere surface complexation on mineral surfaces, P adsorption envelopes and isotherms on aluminum oxide, and iron oxide surfaces were predicted well, resulting in similar protonation–dissociation constants and surface complexation constants for both solids. Nilsson et al. (1992) have also demonstrated the use of CCM to predict P adsorption behavior on goethite surfaces with the aid of the computer model FITEQL. The model was able to predict pH‐dependent inner‐sphere adsorption mechanisms that are represented by three equilibria and intrinsic constants of protonated and nonprotonated surface species. Yao and Millero (1996) used the triple layer model to predict P adsorption on birnessite surfaces. Whereas, a consideration of only inner‐sphere surface species significantly overestimated the total P adsorption at pH < 7, a model with outer‐sphere surface species predicted the experimental data in 0.7‐M NaCl solutions fairly well. Results of equilibrium data modeled using the SCM must be carefully interpreted because the actual adsorption mechanisms is not proven, high degrees of freedom in the adjustable parameters allow one to describe material balance data very well, and the majority of the SCM do not include surface precipitation and other nonadsorption phenomena as part of the model description and prediction (Sparks, 1995b).

B. ELECTROPHORETIC MOBILITY MEASUREMENT STUDIES Electrophoretic mobility (EM) measurements are a useful microscopic approach for not only determining the isoelectric point (IEP) of pure components but also for obtaining information that can be used to indirectly distinguish bulk surface complexes at colloidal–water interfaces. Nonspecific ion adsorption of indiVerent electrolyte at the outside of the shear plane (i.e., formation of outer‐sphere complexes via van der Waals forces) generally does not aVect the IEP but it could cause shifts in the value of EM if the electrolyte is present at high concentration (Hunter, 1981). The shear plane is at the outer edge of the inner part of the double layer and near the outer Helmholtz plane or the Stern layer, depending on the models used to describe the interface (Hunter, 1981). Inner‐sphere complexes, however, cause shifts in both EM and IEP due to specific ion adsorption inside the shear plane (Hunter, 1981). In other words, oxyanion inner‐sphere adsorption does increase the net negative charge on the surface. In some cases, however, inner‐sphere adsorption does not cause shifts in EM (Hunter, 1981). With this knowledge, one can indirectly distinguish the predominant surface complexes on pure colloidal materials using EM measurements.

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There are some EM data suggesting predominant inner‐sphere P complexes on metal oxides. Microelectrophoretic mobility measurements of P adsorbed on ferrihydrite, goethite, and boehmite (g‐AlOOH) have shown that inner‐sphere complexes form due to charge reversal and a lower IEP with increasing P loading level (Anderson and Malotky, 1979; Arai and Sparks, 2001; Bleam et al., 1991; Hansmann and Anderson, 1985; Tejedor‐Tejedor and Anderson, 1990).

C. EX SITU SPECTROSCOPIC STUDIES Ex situ spectroscopic techniques have been extensively utilized to directly distinguish the adsorption mechanisms of P on clays and clay minerals. Parfitt et al. (1975) investigated P adsorption complexation on iron oxides (ferrihydrite, goethite, lepidocrocite, and hematite) using ex situ infrared (IR) spectroscopy. The IR spectra of P adsorbed on iron oxides showed the replacement of two singly coordinated surface hydroxyls, suggesting the formation of bidentate binuclear species. Using the same IR technique, Atkinson et al. (1974) compared the v3 bands of P adsorption complexes on goethite and other model systems [e.g., Co(III)‐P solution complexes]. They suggested that the shift in v3 bands of the adsorption complex was due to binuclear bidentate species (Atkinson et al., 1974). Nanzyo and Watanabe (1982) utilized diVuse reflectance FTIR (DRIFTIR) spectroscopy to investigate P adsorption complexes on goethite over a wide pH range of 3.3–11.9. Goethite background subtracted IR spectra showed that: (1) the surface complexes at the same pH did not change with increasing loading levels up to 197 mole g1 and (2) bidentate bridging complexes were present throughout all pH values based on the IR bands assigned by Parfitt et al. (1975). DRIFTIR spectroscopy was also utilized to re‐investigate P adsorption surface complexation on goethite (Persson et al., 1996). Using the symmetry rule arguments of P v3 bands, Persson suggested that the monodentate surface complex predominantly formed between pH 3 and 12.8, but at intermediate pH, a bidentate complex could not be ruled out. Martine and Smart (1987) utilized X‐ray photoelectron spectroscopy (XPS) to investigate phosphate adsorption complexation on goethite at pH 3 and 12, at high initial P concentrations ([P]i ¼ 5 and 50 mM). They concluded that the replacement of two A‐type surface hydroxyls via a ligand exchange reaction suggested bidentate binuclear adsorption. While ex situ spectroscopic studies have strongly suggested specific P adsorption (inner‐sphere complexes) on Fe and Al oxides, the results have been questioned by many researchers because of the creation of artifacts (i.e., a structural alteration due to vacuum pressure and/or formation of bulk precipitates by drying residual adsorbates) under dry and severely evacuated

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sample conditions. Colloidal–water interfaces in natural settings are usually under near atmospheric pressure. Analyzing samples near environmental conditions (in situ) is more appropriate to obtain accurate information on environmental samples (Goldberg and Sposito, 1985). Accordingly, in the past decade, with the development of in situ molecular scale techniques such as XAS and ATR‐FTIR, one can directly determine oxyanion reaction mechanisms on minerals under environmental conditions that are representative of field settings.

D. IN SITU SPECTROSCOPIC STUDIES 31

P solid‐state magic angle spinning nuclear magnetic resonance (NMR) spectroscopy coupled with the constant‐capacitance model was utilized to investigate the hydrolysis of adsorbed P molecules on boehmite (g‐AlOOH) (Bleam et al., 1991). An inner‐sphere complex was suggested between the pH range of 4 and 11, and a fully deprotonated P surface complex was also reported at pH > 9 (Bleam et al., 1991). Parfitt and Atkinson (1976) combined in situ IR spectroscopy and potentiometric titration experiments to examine P adsorption complexation at the goethite–water interface at pH 3.6 or 5.1. They suggested that binuclear bidentate complexes were predominantly formed at both pH values. Tejedor‐Tejedor and Anderson (1990) used in situ cylindrical internal reflection‐FTIR (CIR‐ FTIR) to investigate P adsorption complexation on goethite between pH 4 and 8. Comparing the spectra of ferric phosphate solutions with those of adsorption complexes on goethite, the formation of protonated and non‐protonated bidentate binuclear and nonprotonated monodentate mononuclear complexes was suggested (Tejedor‐Tejedor and Anderson, 1990). Luengo et al. (2006) have used ATR‐FTIR to elucidate the changes in P surface speciation on goethite surfaces with time (5–400 min). They reported that non‐protonated and protonated bidentate species coexist at pH 4.5, and these species form over time rather than independently. Nonprotonated bidentate species are predominantly formed at pH 7.5 and 9 with an extra unidentified species at low concentration. The formation of these species was not dependent on reaction times. Arai and Sparks (2001) investigated P surface species (reaction time up to 24 h) at the ferrihydrite–liquid (H2O and D2O) interface using ATR‐FTIR spectroscopy. The IR study was combined with peak deconvolution processes to understand the reduction of PO4 Td symmetry that is caused by surface complexation and/or protonation on surface species. They suggested that inner‐sphere surface complexes were nonprotonated bidentate binuclear species (Fe2PO4) at pH > 7.5 and the surface complexes might coexit with diVerent surface species (e.g., monodentate mononuclear) and/or

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A pH 4, 1 year

pH 7.5, 1 year 1025

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978

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1029

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976 1091

1200

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1125 1050

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Wavenumber

900

(cm−1)

Figure 2 ATR‐FTIR spectra of P adsorbed on ferrihydrite at pH 4 (A) and pH 7.5 (B) as a function of residence time. The raw spectra are shown with solid lines, the deconvoluted peaks with dotted lines, and the fitted curve with open circles.

Fe2PO4Na at pH 7.5. The exact identity of the protonated P inner‐ sphere surface complexes (i.e., protonated monodentate mononuclear complexes and/or protonated bidentate binuclear complexes) forming at pH < 7.5 could not be elucidated due to limitations in mid‐IR range FTIR analysis. Arai and Sparks carried out the similar experiments up to 1 year to observe the changes in P surface speciation at the ferriydrite–water interface. ATR‐FTIR spectra of aged P‐reacted ferrihydrite with Gaussian profile fits are shown in Fig. 2A and B. In pH 4 and 7.5 samples (Fig. 2A and B,

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respectively), the fitted Gaussian peaks show triply nondegenerate v3 vibrations except for 1.5 months and 1 year samples at pH 7.5, suggesting that the symmetry of the surface species is C2v and/or lower symmetry. The peak positions of the v3 vibrations are similar to those observed in the previous short‐term adsorption study (Arai and Sparks, 2001), indicating that similar surface complexes might be forming in aged samples. As we previously reported, deuterium exchange did not cause any significant peak shift at pH 7.5, but some shifts occurred at pH 4 under similar loading levels 2.42 mmol m2, suggesting nonprotonated inner‐sphere surface species for pH 7.5 samples and protonated inner‐sphere surface species for pH 4 samples (Arai and Sparks, 2001). Determination of exact bonding environments for pH 4 samples was diYcult based on the FTIR information alone because several diVerent molecular configurations (e.g., monoprotonated bidentate binuclear and/or bidentate binuclear with hydrogen bonding to mineral surfaces) could satisfy C2v and/or lower symmetry with proton associations. Therefore, we only suggested the formation of protonated inner‐sphere complexes at pH 4 in aged samples. While triply nondegenerated v3 vibrations are observed in diVerent aged samples at pH 4, Gaussian peak fit analyses revealed the presence of an additional fourth v3 band (979 cm1) in 1.5 months aged pH 7.5 samples (Fig. 2B). We interpret the fourth peak as the presence of secondary surface species in addition to the predominant surface species observed in the 2‐day‐ aged sample (a bottom spectrum in Fig. 2B). The peak at 978 cm1 is probably one of the v3 vibrations arising from the secondary complexes. It is diYcult to understand whether the secondary surface complexes give two or three v3 vibrations since the strong signals (i.e., three v3 vibrations) from the predominant surface complexes in the 1.5‐month‐aged samples probably overlap with the other v3 vibrations from the secondary complexes. At pH 7.5, ferrihydrite surfaces are predominantly occupied by non‐protonated bidentate bridging PO4 surface species (Arai and Sparks, 2001), and P‐sorbed/ unreacted ferrihydrite surface sites could possibly be accessible by additional monoprotonated (HPO2 4 ) aqueous species to form a minor fraction of: (1) ions and/or (2) inner‐sphere complexes. In fact, one of the v3 diVused HPO2 4 vibrations in the monoprotonated (HPO2 4 ) aqueous species is exhibited around 989 cm1 (Arai and Sparks, 2001), and diVused HPO2 4 molecules in ferrihydrite particles could possibly result in a slight peak shift, producing the forth peak position at 978 and 979 cm1 in 1.5‐month‐aged samples as observed in Fig. 2B. However, we cannot exclude the possibility of secondary inner‐sphere complexes forming. Hesterberg et al. (1999) have used fluorescence yield P K‐edge X‐ray absorption near edge structure spectroscopy (XANES) to investigate solid‐ state P speciation in North Carolina agricultural soils. P K‐edge XANES spectra for Fe‐phosphates were characterized by a unique pre‐edge feature

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near 3 eV (relative energy) that increased in intensity with increasing mineral crystallinity and was very weak for phosphate adsorbed on goethite and aluminum phosphate minerals. Due to spectra resemblances of distinct postedge features, the authors suggested that Ca‐phosphate minerals such as apatite might be forming in the soils. Sato et al. (2005) have used a similar approach to understand the P speciation in silt‐clay fractions of high‐P soils from Southern New York. On the basis of the spectral similarities between unknown samples and Ca and Fe reference compounds [e.g., CaHPO4, CaHPO4 2H2O, Ca3(PO4)2, Ca5(PO4)3OH, and FePO4 2H2O], they suggested that: (1) unamended forest soils contain iron‐associated P species, (2) soils with short‐term manure‐ amended soils contain both iron and Ca‐associated P species, and (3) long‐ term manure‐amended soils contain predominantly Ca‐associated P. Beauchemin et al. (2003) have combined linear combination (LC) XANES fit analyses with principal component analyses (PCA) to better understand P speciation in Canadian soils which had slightly acidic and alkaline soil pH values (5.5–7.6) (Beauchemin et al., 2003). They found sorbed P on Fe and Al oxides in all the soils. While the presence of hydroxylapatite was suggested in two slightly alkaline pH soils, Ca‐associated P species were predominant in an acidic soil. They also reported amorphous iron phosphate minerals in acidic soils of the A horizon. While these studies (Beauchemin et al., 2003; Hesterberg et al., 1999; Sato et al., 2005) showed a useful application of P K‐edge XANES techniques to understand P speciation in bulk soils and soil silt‐clay fractions, the data might have been misinterpreted due to self‐absorption eVects on fluorescence yield XANES spectra of reference compounds. To eliminate the self absorption eVect of concentrated P materials, Toor et al. (2006) have demonstrated the use of total electron yield XANES data to appropriately understand the P speciation in biosolids (Toor et al., 2006). The direct P speciation of soils using LC XANES fit and PCA analyses must be carefully performed since the results can be misinterpreted due to limitations using these analyses. Beauchem et al. (2003) pointed out that the LC XANES analyses of P K‐edge XANES data can be inherently restricted by: (1) the data quality and (2) correct choice of appropriate standards. Furthermore, several researchers have indicated that PCA of X‐ray absorption spectra are not sensitive enough to pick up subtle diVerences in spectral features (e.g., diVerences in ligand coordination) (Arai et al., 2006; Beauchemin et al., 2003). Therefore, the number of significant components suggested by indicator functions must be only used as a guide rather than drawing conclusions. In more recent studies, several researchers have used P K‐edge XANES techniques to understand P adsorption mechanisms at the metal–oxyhydroxide interface. Khare et al. (2005) investigated the P adsorption mechanisms in binary adsorbent systems (e.g., mixture of iron oxyhydroxides and aluminum

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oxide) using P K‐edge XANES analyses. Changes in the full width at half‐ maximum height in the white line peak were used to distinguish diVerent binding mechanisms. While P precipitation was observed in an Al oxide single‐mineral system as well as a goethite/bohemite binary system, only adsorption was observed in ferrihydrite and goethite single‐mineral systems. Arai and Sparks recently conducted fluorescence yield P K‐edge XANES measurements on diVerent P salts, minerals, and P‐reacted ferrihydrite (suspension density, 1.25 g liter1; I, 0.1‐M NaCl; and initial P concentrations, 1 mM). In Fig. 3A–D, florescence yield XANES spectra of P minerals/salts, P‐reacted ferrihydrite, and PO4(aq) samples are shown. P K‐edge XANES spectra of reference iron phosphate minerals, strengite (FeIII(PO4)2H2O), II II III barbosalite (FeII FeIII 2 ðPO4 Þ2 ðOHÞ2 ), and rockbridgeite (ðFe ; Mn ÞFe4 ðPO4 Þ3 ðOHÞ5 ) are compared in Fig. 3A. Strong white line peaks can be observed in all spectra as the result of a P 1s electron transition into an unoccupied valence state of PO4 sp3 hybridized orbitals. Distinctive postedge resonance features, at 5–25 eV relative energy, are similar in crystalline and amorphous strengite but barbosalite and rockbridgeite show diVerent features. The relative energy regions of pre‐edge doublet features (indicated by dotted lines between 6 and 3 eV) are similar in strengite, barbosalite, and rockbridgeite. The pre‐edge region of P K‐edge XANES spectra includes a sharp white line peak resulting from electronic transitions of the core electronic states in the conduction band (Behrens, 1992; Fendorf and Sparks, 1996). The electronegativity, number of nearest neighbors, and coordination environment (i.e., molecular symmetry) of the absorbing atoms could influence the intensity and the position of the pre‐edge features (Behrens, 1992; Behrens et al., 1991; Wong et al., 1984). On the basis of a P XANES study on several transition metal phosphates by Okude et al. (1999), the pre‐edge features in an Fe(III)‐P salt [i.e., Fe4(PO4)3(OH)3] were attributed to interactions between P 3p states [i.e., sp3 hybridization of the phosphate (PO4) tetrahedral molecule] and 3d5 electronic state of Fe(III). It is also important to note that the pre‐edge feature is not present when there is no direct interaction between P tetrahedra and Fe(III) octahedra [i.e., HPO2 4 (aq) spectrum in Fig. 3A]. The position and intensity of the pre‐edge features also depend on the number of d‐electrons in the transition metal associated with phosphate (Okude et al., 1999). The eVects of d0–d8 electronic state of metal phosphate salts on the P XANES pre‐edge features were re‐investigated after the study by Okude et al. (1999), and some of them are reproduced, along with new results (Fig. 3B). As previously reported by Okude et al. (1999), the pre‐edge intensity decreases with increasing number of d‐electrons, and the position (indicated by solid lines in Fig. 3B) shifts to a lower energy, with a gap between d5 and d6 electronic states (Fig. 3B). Moreover, the peak maxima of the white line shift to higher energy with decreasing number of d‐electrons. Although vivianite [i.e., a predominantly d6 Fe(II)‐based P mineral] shows a subtle pre‐edge

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A

B

Normalized intensity

Normalized intensity

d0, CaHPO4 2− HPO4 (aq)

Rockbrigeite Barbosalite

d4, CrPO4 d5, Strengite (FeIIIPO4·2H2O) d6, Vivianite (Fe3II[PO4]2·8H2O) d7, Co3(PO4)2

Crystalline strengite

d8, Ni3(PO4)2

Amorphous strengite

d10, Zn3(PO4)2 −10

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10 20 30 40 Relative energy (eV)

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−10

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Figure 3 (A) P K‐edge XANES spectra of synthetic amorphous strengite (FeIIIPO4 2H2O), II crystalline strengite (FeIIIPO4 2H2O), barbosalite (FeII FeIII 2 ðOHÞ2 ðPO4 Þ2 ), rockbrigite ((Fe , II III 2 Mn )Fe 4 (PO4)3(OH)5), and HPO4 (aq). Relative energies are presented with respect to the absorption edge energy position of ferric phosphate minerals, (B) P K‐edge XANES spectra of transition metal phosphate salts; dx (x, number of d electrons of the predominant transition metal). Solid lines and dotted lines are pre‐edge and white line peak maxima, respectively (modified after Okude et al., 1999), (C) P K‐edge XANES spectra of P‐reacted ferrihydrite at pH 4.0 as a function of residence time, and (D) P K‐edge XANES spectra of P‐reacted ferrihydrite at pH 7.5 as a function of residence time.

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feature at a similar position (2150 eV) due to a small amount of d5 Fe(III) impurity, it does not contain a strong pre‐edge feature that is observed in amorphous strengite [i.e., d5 Fe(III)‐P mineral]. In a similar tetrahedral oxoanion (i.e., sulfate) XANES study, Myneni (2000) compared the energy position of the pre‐edge feature in aqueous sulfate and Fe(III)‐SO4 salts [i.e., coquimbite, (Fe,Al)2(SO4)3 9H2O; copiapite, MgFe4(OH)2(SO4)6 10H2O; jarosite, KFe3(SO4)2(OH)6; ferric ammonium sulfate, FeNH4(SO4)2 12H2O] that contain diVerent SO4‐Fe(III) coordination environments (bidentate binuclear linkages in coquimbite and copiapite, tridentate linkages in jarosite, and hydrogen bonds in ferric ammonium sulfate). These data indicated that the energy position of the pre‐edge feature in these Fe(III)‐SO4 salts depends on the number of Fe(III) polyhedra connected to each sulfate polyhedron surface, and the pre‐edge feature that was absent in the samples had no direct interaction between Fe(III) polyhedra and sulfate (i.e., ferric ammonium sulfate and aqueous sulfate) (Myneni, 2000). In the case of Fe(III)‐P salts/minerals, the relationships between specific Fe(III)‐O‐P coordination environments and pre‐edge features have not been well investigated. Therefore, we further investigated the pre‐edge feature characteristics (i.e., shape and energy position) in natural/synthetic Fe(III)‐P minerals (barbosalite, rockbridgeite, and strengite) containing diVerent P tetrahedral attachment on iron octahedra (e.g., bidentate binuclear configuration) and aqueous phosphate (i.e., HPO2 4 ). Phosphate tetrahedra and Fe(III) octahedra coordination environments in these minerals were previously studied by Rose et al. (1997). They described the coordination environments using a notation (NXMY) based on the number of corners of the P tetrahedron in association with the Fe(III) octahedron and the number of Fe atoms in the second coordination sphere. The notation NXMY was used to describe the diVerent linkages: N is the number of corners of the P tetrahedron involved in the linkage, X is the type of linkage between the polyhedra (i.e., C for corner and E for edge), M stands for the number of iron atoms in the second coordination sphere of P (for one linkage), and Y is the type of linkage between iron octahedra (C for corner and F for face), when more than one Fe(III) octahedra is in the second coordination sphere (Rose et al., 1997). The linkages are 1C1 (monodentate mononuclear corner attachment on an iron octahedron) for strengite, 1C2F (monodentate binuclear corner attachment on the center of the face sharing two iron octahedra) for both barbosalite and rockbridgeite, and 2C2C (bidentate binuclear corner attachment on two corners sharing iron octahedra) for both barbosalite and rockbridgeite (Rose et al., 1997). While the aqueous phosphate spectrum shows no pre‐edge feature [i.e., no Fe(III)‐O‐P linkages], the strong doublet pre‐edge features are consistently present at a similar energy range (i.e., 3 to 6 eV) in all spectra regardless of the diVerent types of Fe(III)‐O‐P linkages (i.e., 1C1, 1C2F, 2C2C) (Fig. 3A).

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Although the shape of the pre‐edge feature is slightly varied in these Fe(III)‐P minerals, the presence of direct P‐O‐Fe(III) linkages is indicated in these minerals. Unlike the Fe(III)‐SO4 reference salts used in Myneni’s S XANES study (Myneni, 2000), our natural Fe(III)‐P minerals contain: (1) transition metal impurities [i.e., d6 Fe(II) and d5 Mn(II)] coordinated with PO4 which weaken the pre‐edge feature seen in Fe(III)‐P compounds and (2) mixed PO4‐Fe(III) linkages in barbosalite and rockbrigite (i.e., 2C2C and 1C2F). Therefore, it is diYcult to conclude that the shape, energy position, and intensity of the pre‐edge features are strictly attributed to the specific coordination environments between P tetrahedra and Fe(III) octahedra. Although observed pre‐edge features do not provide clear evidence for distinguishing monodentate mononuclear from bidentate binuclear linkages, it is safe to say that the pre‐edge features observed in the Fe(III)‐P solids arise from a direct interaction between P tetrahedra and Fe(III) octahedra, and this feature can be used as a signature of covalent bonds between PO4 and iron(III) octahedra. Myneni previously used this feature as an indication of inner‐sphere SO4 coordination environments on goethite surfaces (Myneni, 2000). Electrophoretic mobility measurements are often used to distinguish diVerent complexation mechanisms (i.e., inner‐sphere and outer‐ sphere complexes). However, charge reversal in the presence of oxyanions is diYcult to interpret since formation of ternary complexes with inert background electrolyte ions and/or formation of surface precipitates could mask the experimental results. The pre‐edge features observed in PO4/SO4 K‐edge XANES measurements can be useful for identifying the inner‐sphere P/S oxyanion coordination environments on the Fe(III) oxide surfaces. In Fig. 3C and D, XANES spectra of adsorption kinetic samples at pH 4 and 2 7.5 and H2 PO 4 (aq) and HPO4 (aq) samples are shown. The distinctive postedge resonance features seen in the Fe(III)‐P reference minerals at 0–30 eV (Fig. 3A) are absent in all of the sorption kinetic samples (Fig. 3C and D), indicating no predominant fraction of Fe(III)‐P precipitates. This might be due to deconstructive interferences of the outgoing multiple scattered photoelectron wave function from predominant two‐dimensional adsorption complexes which overwhelm weak resonance features of Fe(III)‐P precipitates, if any form. Several researchers have previously suggested the formation of iron phosphate precipitates at mineral surfaces (i.e., ferrihydrite, goethite, and hematite) on P sorption using various ex situ spectroscopic techniques (i.e., Auger, DRIFT, XRD, XPS, SIMS, and TEM) (Johansson et al., 1998; Martine and Smart, 1987; Martine et al., 1988; McCammon and Burn, 1980; Nanzyo, 1986). However, there was no evidence for the formation of ferric phosphate precipitates under our in situ experimental reaction conditions. Saturation indices (a maximum 0.932) estimated using the equilibrium constant of synthetic strengite (Nriagu, 1972) and dissolved total Fe concentrations also indicate that all systems were undersaturated with respect to synthetic strengite.

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Compared to the aqueous PO4(aq) spectra, weak pre‐edge/shoulder features (indicated by two dotted lines between 6 and 3 eV) appear after 5 min in the P adsorption samples at both pH values (Fig. 3C and D). It seems that the pre‐edge features become more pronounced with increasing reaction times at both pH values (indicated by arrows in Fig. 3C and D). These peak positions are similar to those in strengite, indicating the formation of inner‐ sphere P tetrahedral linkages on the Fe(III) octahedral structures. This is in good agreement with previous results of ATR‐FTIR analyses and electrophoretic mobility measurements (Arai and Sparks, 2001). In Fig. 4, the first derivative of selected spectra from Fig. 3A–C are summarized. In this figure, one can clearly see that the pre‐edge features in the adsorption samples become more pronounced with increasing aging time from 1.5 days to >11 months at both pH values, and the features in aged adsorption samples (indicated by open circles) become near doublets which are also observed in amorphous and/or crystalline strengite, barbosalite, and rockbrigeite (indicated by an arrow). This suggests that similar inner‐sphere P‐O‐Fe(III) linkages are present in 11‐ months‐aged adsorption samples at pH 4 and 7.5. The pre‐edge features intensify with aging from 1.5 days to 11 months at both pH values. Since XANES spectra are normalized with respect to per P atom, the intensified

Crystalline strengite

First derivative intensity

Amorphous strengite pH 4, 11 months pH 4, 1.5 days pH 4, 1 h pH 7.5, 12 months pH 7.5, 1.5 days pH 7.5, 1 h

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50

Relative energy (eV) Figure 4 First derivative of selected P K‐edge XANES spectra from Fig. 3A, C, and D.

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features cannot be explained by an increase in P loading levels. An increase in the amount of inner‐sphere P‐O‐Fe(III) linkages seems to be the only explanation to support the pronounced pre‐edge feature. Interestingly, the P loading levels remained nearly equal (2.73 0.03 mmol m2) after 1.5 days in the pH 4 system. The lack of a pre‐edge feature in the 1.5‐day sample compared to the 11‐month sample is probably attributed to the presence of secondary species (e.g., diVused ions), and these complexes were gradually converted into more stable inner‐sphere species after 11 months. Similar reaction mechanisms can be also suggested in the pH 7.5 system. The intensified pre‐edge features are due to changes in surface speciation [i.e., an increase in inner‐sphere P‐O‐Fe(III) linkages]. Two possible mechanisms are postulated to explain the pre‐edge feature in aged samples. First, an increase in inner‐sphere Fe(III)‐O‐P linkages can be supported by the formation of PO4‐bridged ferric polymer complexes. The diVused P ions within ferrihydrite aggregates could possibly react with unreacted ferrihydrite particles to form PO4‐bridged ferrihydrite polymers. In a similar tetrahedral oxianion [i.e., arsenate (As(V))] adsorption study on ferrihydrite, Waychunas et al. (1993) observed large As(V)‐Fe coordination numbers (i.e., >3.02) in As(V) sorbed ferrihydrite samples based on As and Fe K‐edge EXAFS analyses. They suggested that ferrihydrite might contain substantially smaller basic crystalline units that are freely accessible by disordered aggregates as well as dissolved arsenate anions as long as local geometry permits (Waychunas et al., 1993). Since AsO4 is often used as an analogue of PO4 due to similar chemical properties (e.g., solution speciation with respect to protonation constants), it is possible that PO4 molecules could induce the formation of ferric polymer chains. Since our experimental pH values (4 and 7.5) are below the PZSE of ferrihydrite (8), the surface charge density of ferrihydrite particles not reacted with P is positively charged. These positively charged particles could be electrostatically attracted to P‐sorbed ferrihydrite surfaces which are negatively charged due to specific P adsorption. Previous electrophoretic mobility measurements have shown that charge reversal can occur via inner‐sphere P surface complexation on ferrihydrite (Arai and Sparks, 2001). The particle interaction could eventually lead to the formation of multi‐Fe(III) polymer layers/clusters, entrapping/bridging P adsorption complexes within the structure. Interestingly, Anderson et al. (1985) earlier suggested a similar PO4‐ partitioning mechanism on Fe oxyhydroxide (i.e., goethite) based on XRD/ electron microscopic experimental data (Anderson et al., 1985). An aggregate order of goethite particles significantly increased after goethite particles were reacted with phosphate, and isotopically exchangeable P was decreased with increasing P uptake (Anderson et al., 1985). In aged P‐reacted ferrihydrite surfaces, similar P‐partitioning mechanisms might have been occurred. Second, ferric phosphate surface precipitation mechanisms on the ferrihydrite surface could also result in an enhancement of pre‐edge features

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[i.e., an increase in Fe(III)‐PO4 linkages]. Ler and Stanforth (2003) reported that the Zeta (x) potential of P sorbed on a goethite surface at pH 5 with time reverted to that observed for goethite surfaces without P (Ler and Stanforth, 2003). The long‐term reactions most likely involve the dissolution of ferric ions from goethite surfaces and the subsequent reaction between ferric (aq) and surface‐bound P. In our P–ferrihydrite system, it is possible that dissolved ferric ions are interacting with diVused P ions/surface‐bound P, resulting in the formation of ferric phosphate surface precipitates on ferrihydrite surfaces with increasing time.

V. RESIDENCE TIME EFFECTS ON PHOSPHATE ADSORPTION AND DESORPTION IN SOILS AND SOIL COMPONENTS A. RESIDENCE TIME EFFECTS THEORY In environmental settings, it is important to consider the residence time (aging) eVect on contaminant bioavailability, transport, and remediation. Soils and sediments are nearly always at disequilibrium with respect to ion transformations (Sparks, 1987). The rate of bioavailability can be reduced or increased in the environment with increasing time (Pignatello and Xing, 1995). The residence time eVect has also been described by irreversible, hysteretic, and nonsingular reactions, and can be suggested by the observation of two slow reaction processes: (1) a continuous slow adsorption of the adsorptive onto the adsorbent with increasing time and (2) a slow desorption of the adsorbate from the adsorbent with increases in time. A slow adsorption process usually occurs after an establishment of quasi‐ equilibrium following an initial rapid reaction for a few hours. In thermodynamics, the definition of slow desorption is that it is diYcult for the desorptive to overcome the total activation energy created during adsorption. In other words, the activation energy of desorption is less than the sum of the activation energy for adsorption and the energy formed during adsorption [Edesorption Eadsorption þ DH, where Edesorption ¼ activation energy for desorption, Eadsorption ¼ activation energy for adsorption (0), and DH ¼ energy formed during adsorption]. Formation of inner‐sphere complexes via ligand exchange and/or the transformation of amorphous to crystalline materials generate a greater DH, increasing the value of the total activation energy. Slow desorption phenomena are attributed not only to chemical factors (e.g., ligand exchange and chemisorption) but also to physical factors (diVusion). Ions trapped in mesopores (interparticle) between aggregates

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and micropores (intraparticles) within an individual particle fissure are diYcult to extract and/or desorb with any desorptive.

B. SLOW ADSORPTION Continuous slow adsorption of P on soils and soil components [i.e., amorphous Al(OH)3, natural allophane, ferrihydrite, hematite, goethite, a‐Al2O3, and kaolinite] over diVerent timescales (hours to months) was reported by many researchers (Anderson et al., 1976; Barro´n and Torrent, 1995; Barrow, 1974, 1985; Beek and van Riemsdijk, 1982; Black, 1942; Colemann, 1944; Edzwald et al., 1976; Fuller et al., 1993; Hingston et al., 1974; Hsu and Rennie, 1962; Kafkafi et al., 1967; Madrid and Posner, 1979; Okajima et al., 1983; Parfitt, 1979, 1989; Ryden and Syers, 1977; van Riemsdijk et al., 1977; Willett et al., 1988). While some studies documented continuous slow P adsorption on phyllosilicate and aluminum and iron oxyhydroxides over timescales of hours, others carried out the experiments up to days–months on soils and soil components. Edzwald et al. (1976) observed rapid P adsorption on kaolinite, montmorillonite, and illite at pH 7–8 during the initial 4 h, and this was followed by a slow continuous uptake after 24–72 h (Edzwald et al., 1976). Madrid and Posner (1979) also documented slow P uptake on goethite after 1–24 h at pH 4.25–10.25. Ryden and Syers (Ryden and Syers, 1977; Ryden et al., 1973, 1977a) studied long‐term P adsorption on natural and synthetic goethite, hydrous ferric oxides and four New Zealand soils (Egmontblack loam, Porirua fine sandy loam, Okaihau gravelly clay, and Waikakahi silit loam). In all adsorbents, an initial fast reaction was completed within 20 h, followed by slow P adsorption after 192 h. Hsu and Rennie (1962) investigated P sorption on amorphous aluminum hydroxides at pH 3.8, 4.2, and 7. While short‐term experiments showed only an initial rapid adsorption within 24 h at pH 4.2 and 7, continuous slow adsorption was observed after 528 h at pH 3.8 in the long‐term experiments. Black (1942) carried out P adsorption experiments on a Cecil clay for days. The slow uptake continued from 48 h to 30 days over a wide range of pH values (2.5 to 8). An increase in P adsorption with time at acidic pH values was much greater than at alkaline pH values. van Riemsijk et al. (1977) investigated P adsorption on aluminum oxide and a‐Al2O3 at pH 5 and 6 using inorganic synthetic sewage water. In both solids, an initial fast adsorption was completed within 1–4 days, and this was followed by a slow adsorption up to 70 days. Haseman et al. (1950) studied long‐term P sorption on clay minerals (montmorillonite, illite, and kaolinite) at pH 3–7. While slow adsorption was observed for all clay minerals up to 200–300 h, long‐term experiments on gibbsite and goethite showed a slow continuous

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uptake even after 1000 h at pH 3–5. Willet et al. (1988) have documented long‐term slow P sorption on ferrihydrite aggregates after 90 days at pH 4. The P migration into the internal adsorption sites of ferrihydrite aggregates was evident in electron microprobe analyses. Arai and Sparks have recently investigated the eVect of initial P concentrations (0.6–2.6 mM) on P adsorption kinetics on ferrihydrite surfaces (suspension density, 0.5 g liter1; I, 0.1‐M NaCl and pH 4 and 7.5). They observed fast adsorption within the initial 4 h under all reaction conditions (insets of Fig. 5A and B), and the rapid adsorption was followed by slow continuous adsorption after 650 h. At steady state (670 h), total P adsorption was consistently greater in the pH 4 systems than in the pH 7.5 systems at respective initial P concentrations (Fig. 5A and B).

C. SLOW DESORPTION PROCESS

AND

HYSTERESIS

Many researchers utilizing batch techniques (e.g., replenishment techniques) to observe short‐term (<24 h) P desorption processes from soils and soil components have reported that the process was often biphasic (a fast initial desorption followed by slow desorption). Two studies examined P desorption (>120 h) on synthetic goethite and ferrihydrite using batch replenishment techniques (diluting the reacted samples at a constant I ) (Madrid and Posner, 1979; Ryden and Syers, 1977). Desorption phenomena were biphasic, a fast reaction for a few hours was followed by a slow reaction for over 100 h (Madrid and Posner, 1979; Ryden and Syers, 1977). Biphasic P desorption phenomena were also observed for kaolinite using a similar batch technique (Bar‐Yosef and Kafkafi, 1978). Overall, the desorption process was divided into a rapid and a slow first‐order reaction, and the rate coeYcient for the rapid reaction was about four times higher than the slower reaction. The traditional batch technique is not usually suitable to investigate desorption processes due to: (1) possible readsorption of reaction products due to their accumulation in the bulk solution, (2) pH fluctuation during desorption, and (3) establishment of a diVerent equilibrium after each replenishment step. Flow systems and batch techniques (e.g., an ion exchange resin) that employ a ‘‘sink’’ for the desorbed species prevent the build up of reaction products in the bulk solution, and are preferable kinetic methods for studying apparent desorption behavior for long‐time periods (>24 h). Phosphate release from P‐reacted iron oxyhydroxide (i.e., ferrihydrite, goethite, and hematite) coated silica was investigated using a flow‐through method (Freese et al., 1999). After 4400 min of desorption at pH 4, the relative amount of desorbed P ¼ Pdesorb/Psorb was found to be of the order of ferrihydrite < goethite < hematite.

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A 120,000 100,000

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Time (h) Figure 5 P adsorption kinetics (suspension density, 0.5 g liter1 and I, 0.1‐M NaCl) at the ferrihydrite–water interface at (A) pH 4 and (B) pH 7.5 as a function of initial P concentrations (0.6, 1.3, and 2.6 mM). The adsorption reactions up to 24 h are magnified in insets.

Several kinetic models have been used to successfully describe biphasic P desorption processes from soils and soil components. These include zero‐ order (Onken and Matheson, 1982), first‐order (GriYn and Jurinak, 1973), second‐order (Kuo and Lotse, 1973), third‐order (Onken and Matheson, 1982),

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parabolic diVusion (Evans and Jurinak, 1976), and Elovich (Atkinson et al., 1970; Chen, 1977; Chen and Clayton, 1980). One of the reasons a single‐ kinetic model is highly applicable to describe biphasic reactions is that experimental conditions (i.e., the length of experimental periods associated with disequilibrium) are often appropriate for the assumptions of the model (Sparks, 1989). Aharoni and Suzin (1982) used heterogeneous diVusion models to describe segments of kinetic processes (initial, intermediate, and long term). These models were approximated by a sequence of parabolic, Elovich, and exponential equations. In some of early P kinetics research, Australian researchers compared P adsorption and desorption processes on soils and soil components and investigated the hysteresis eVect, in which an adsorption isotherm curve does not coincide with a desorption isotherm curve. Using a batch equilibrium study (<24 h), Barrow (1983b) showed the slight shifts between adsorption and desorption isotherm curves on an Australian sandy loam. Ryden and Syers (1977) utilized the same experimental approaches to compare the P hysteresis eVect on soils and ferrihydrite. Desorption isotherm curves on the soils and ferrihydrite were greatly shifted from adsorption isotherm curves, and ferrihydrite showed greater irreversibility than the soils (Ryden and Syers, 1977). The concentration eVect on hysteresis was also investigated. A wide range of initial P ([P]i ¼ 0.001–1 mM) was reacted with kaolinite, and then the desorption was investigated using isotopic exchange (Kafkafi et al., 1967). A hysteresis eVect was observed over all concentration ranges. Incubation time highly influences the reversibility of adsorbed P from soils and soil components. Barrow reported that P desorption from Australian soils was rapid when aging time was short (<24 h), but the desorption process became much slower when aging time was longer (>24 h) (Barrow, 1979a; Barrow and Shaw, 1975). Madrid observed similar irreversible P desorption from goethite (Madrid and Posner, 1979). P desorption decreased with increasing aging time from 1 to 23 h. A similar study was also conducted on a synthetic goethite and Australian clay loam using longer desorption experiments (>96 h) and varying incubation time. The irreversibility was enhanced with increasing aging time (Barrow, 1979a; Madrid and Posner, 1979). The irreversibility could also occur within a short adsorption reaction (<15 s). In our recent investigation, we have observed similar aging eVects on long‐ term (up to 19 months)‐aged P‐sorbed ferrihydrite at pH 4 and 7.5 (suspension density, 1 g liter1; I, 0.1‐M NaCl; and initial P concentrations, 1 mM for the pH 4 system and 0.6 mM for the pH 7.5 system). Two diVerent initial P concentrations were chosen to achieve nearly 100% adsorption within 2 days, so that the loading levels for all aged samples were approximately equal at each pH value (2.7 mmol m2 at pH 4 and 1.62 mmol m2 at pH 7.5) prior to each desorption experiment. It is important to mention that ferrihydrite

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Time (h) Figure 6 Residence time eVects on P desorption kinetics from ferrihydrite at pH 4.0 (A) and pH 7.5 (B). Incubation times of P‐reacted ferrihydrite suspensions prior to desorption experiments are described in legends (suspension density, 1 g liter1; I, 0.1‐M NaCl; and initial P concentrations, 1 mM for the pH 4 system and 0.6 mM for the pH 7.5 system).

transformation into crystalline phases such as goethite and hematite was not observed in 1‐year‐aged samples at both pH values via conventional bulk XRD analyses. Figure 6A and B shows short‐term (24 h) P desorption data at pH 4 and 7.5, respectively. The stirred flow method was used so that the adsorbents are exposed to a greater mass of ions than a static batch system,

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and the flowing solution continuously removes reaction products (i.e., desorbed species) (Sparks, 1989). Biphasic P desorption reactions were observed at both pH values (Fig. 6A and B). At pH 4, the initial fast reaction occurred during the first 3–5 h, and slow reactions followed up to 24 h, and total P desorption decreased from 4.81% to 0.92% with increasing aging time from 2 days to 10 months, suggesting a residence time eVect. Similarly, biphasic desorption behavior was observed at pH 7.5 (Fig. 6B). Total desorbable P decreased from 10.07% to 3.83% with increasing aging time from 2 days to 19 months at pH 7.5. It is interesting that total desorbable P is much less at pH 4 even though the initial loading levels at pH 4 are much greater than those at pH 7.5. The mechanisms responsible for the residence time eVect are not clearly defined; however, it has been ascribed to: (1) solid‐state diVusion, or intra‐ and interparticle diVusion (Bar‐Yosef and Kafkafi, 1978; Barrow, 1983a; Bolan et al., 1985; Cabrea et al., 1981; Enfield et al., 1981; Evans and Jurinak, 1976; Parfitt, 1989; Ryden et al., 1977a; Torrent et al., 1992; van Riemsdijk et al., 1984; Willett et al., 1988), (2) surface precipitation (e.g., Ler and Stanforth, 2003), and (3) higher energy binding on surface structures via chemical reconfiguration (Beek and van Riemsdijk, 1982; Fendorf et al., 1992; Pignatello and Xing, 1995).

D. SOLID‐STATE, INTER‐, AND INTRAPARTICLE DIFFUSION In situ soil chemical reactions between ions and solids are ultimately limited by the process of diVusion (McBride, 1994). For example, if the initial oxyanion reaction on soil materials is completed rapidly, then the unreacted oxyanions build up in solution until the rate of diVusion of oxyanions in the adsorbents equals the overall reaction rate. Fast chemisorption (e.g., the ligand exchange reaction) takes place at external sites where ions in the bulk fluid can easily access the sites. The slow adsorption processes occur at sites within soil particle aggregates, and are often controlled by diVusion processes. Four diVerent diVusion processes have been proposed: film, mesopore (2 nm), micropore (<2 nm), and solid‐state diVusion (Pignatello and Xing, 1995). Film diVusion involves the transport of an ion or molecule through a boundary layer or film (water molecules) that surrounds the particle surface (Sparks, 1995a). Mesopore (interparticle) diVusion takes place between aggregates. Micropore (intraparticles) diVusion involves ion penetration in an individual particle fissure. Solid‐state diVusion involves diVusion in the solid, and it can be a rate‐limiting step for all diVusion processes. The rate of these diVusion processes is dependent on the size and chemical properties of the diVusant and the size, shape, tortuosity, and discontinuity of the particle pores (Pignatello and Xing, 1995).

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One way to investigate the diVusion process is to model slow adsorption processes using models accounting for diVusion reactions. Several modeling studies have indirectly suggested P solid‐state mesopore and micropore diVusion processes in natural materials (Barrow, 1983a; Enfield et al., 1981; van Riemsdijk et al., 1984). In an early P adsorption‐modeling study, Barrow (1983a) developed a mechanistic model to describe slow P adsorption processes on soils. The model had three components: (1) reaction between phosphate ions and variable charge mineral surfaces, (2) assumption that soils consisted of an assemblage of elements, each with diVerent values for the binding constant and/or initial electrostatic potential, and (3) the assumption that the solid‐state diVusion processes are driven by the gradients of electrical and chemical potentials at the adsorbent surface (Manning, 1968; Rickert, 1982). As a result, the model closely described P adsorption processes as influenced by concentration, pH, temperature, and reaction time. It was suggested that long‐term (90 days) P adsorption is predominantly attributed to P penetration into the soil particles. In a later study, Bolan et al. (1985) utilized Barrow’s mechanistic model to describe slow P adsorption processes on soil components (i.e., amorphous iron and aluminum oxides). Models that fit the data well indirectly suggested that slow uptake of P was caused by redistribution of P into the interior of amorphous metal oxide particles via solid‐state diVusion. A similar model successfully described continuous P sorption on a ferric hydroxide gel (Ryden et al., 1977a). While Barrow’s mechanistic model has been extensively used to describe oxyanion adsorption phenomena, other models have also been utilized to investigate slow adsorption processes. While these modeling approaches indirectly suggest a diVusion‐controlled reaction, they do not provide any information on what soil physical properties (e.g., crystallinity of adsorbents) are associated with slow adsorption. Several researchers have pointed out that slow adsorption processes could be aVected by crystallinity and particle morphology. Phosphate adsorption on goethite and lepidocrocite was compared using a batch adsorption technique (Cabrea et al., 1981). Lepidocrocite exhibited more pronounced continuous slow sorption than goethite, and adsorbed P was more irreversible in lepidocrocite. Cabrea et al. (1981) suggested that the enhanced slow reaction and irreversibility were attributed to P micropore diVusion into the less crystalline nature of lepidocrocite particles. Parfitt (1989) also indirectly suggested that P particle diVusion was caused by poor crystallinity and the greater porosity of metal oxides. A comparison of 1‐ and 30‐day P sorption isotherm experiments showed that slow P sorption was observed in only amorphous goethite and ferrihydrite and not in highly crystalline goethite (Parfitt, 1989). Later, Torrent et al. (1992) conducted experiments that supported the above findings. The slow P adsorption on 10 goethite rich soils was described using a modified Frendlich equation including a rate term (Torrent et al., 1992). The study showed that the extent

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of the slow continuous reaction was correlated to the ratio between micropore surface area and total surface area, as well as impurities due to amorphous Fe oxides (Feox). Willet et al. (1988) found: (1) long‐term (>200 days) adsorption and (2) rapid (<24 h) migration of phosphate (mesopore diVusion) into the aggregated particles of ferrihydrite using electron microprobe analyses. Long‐term P adsorption kinetics at pH 4 exhibited an initial fast reaction followed by slow sorption up to 260 days, and a maximum sorption capacity of 1.2 mM g–1. The distribution of P moved toward the core of ferrihydrite aggregate particles as time increased from 1 h to 1 day. Electron microscopic studies have also shown that the crystalline morphologies of goethite are related to P irreversibility (Torrent et al., 1990). After sequential acid (5‐M HCl and 0.5‐M H2SO4)/base (0.1‐M KOH) extractions of adsorbed P from 31 diVerent goethites, the amount of nonextractable P (P was adsorbed at pH 6 for 24 h) was correlated to the observation of solid morphologies [thin, multidomainic laths (V‐shaped interdomainic grooves), and slit‐shaped macropores of multidomainic laths]. DiVusion‐controlled reactions can also be suggested by a low energy of activation (<42 kJ mol1), Ea (Sparks, 1989). The results from a temperature‐ dependent sorption/desorption study can be applied to the Arrhenius equation (k ¼ AeEa =RT ) to obtain Ea (Sparks, 1995a). The Ea of the slow reaction between phosphate and diVerent morphologies of goethite was investigated (Torrent, 1991). The slow sorption was described by a modified Frendlich equation including time and activation energy (Ea) terms. Activation energy values ranged from 38 to 80 kJ mol1, suggesting that the P sorption reaction might involve a penetration of P into the diVerent crystalline matrices.

E. SURFACE PRECIPITATION Whereas an adsorption complex has a two‐dimensional surface structure, surface precipitates involve a three‐dimensional growth of the adsorbate on the adsorptive surface. The magnitude of the precipitation growth is a function of the saturation of the adsorptive in the bulk fluid. When the adsorptive is supersaturated with respect to the precipitate in the bulk fluid, a bulk precipitate forms rapidly. However, when the system is even undersaturated with respect to the bulk precipitate, precipitates could still form. Such a precipitate is called a surface precipitate. Surface precipitates are categorized into one of three groups based on formation mechanisms. These are: (1) polymerization of the adsorptive at the surface, (2) coprecipitation which forms between dissolved coions and the adsorbate in the bulk solid, and (3) precipitates composed of ions from the bulk fluid (e.g., hydrolysis products) (Chisholm‐Brause et al., 1990;

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Farley et al., 1985). Polymerization occurs as the result of continued chemisorption. Newly created sites, by monolayer coverage of the adsorbate, facilitates the formation of the precipitates (multilayer surface coverage). Coprecipitation processes can also be facilitated by an increase in the surface coverage of the adsorbent, and then the precipitation of adsorbate ions form with dissolved coions of the solids (Sparks, 1995b). Several hypotheses have been suggested to describe the formation of surface precipitates. The first theory states that the dielectric constant of the solution near the surface is less than that of the bulk fluid because the mineral surfaces are polar due to the dipole moment of adsorbed water molecules (O’Day et al., 1994). Since the activity of the ions in the bulk fluid is inversely proportional to the dielectric constant of water, the lowered dielectric constant can facilitate the formation of ion activity products. The second theory states that the activity of the solid phase is less than unity (Sposito, 1986). The activity of the pure solid (i.e., FePO4) can be assumed to be 1; however, if a mixed solid (FexAl1xPO4) is formed, the activity is less than 1. FexAl1xPO4 will precipitate prior to FePO4. The following empirical equation describes formation of the surface precipitate, which is favored over the pure solubility products: ðIAPÞi ¼ gi Xi Kiso

ð5Þ

where gi is the activity coeYcient of the solids i, Xi is the mole fraction of the solid, and Kiso is the bulk precipitate of the pure minerals i (van Riemsdijk and Lyklema, 1980). The third theory states that sterically similar sites promote further nucleation (McBride, 1991). Sterically similar sites indicate that the crystal lattice has energy barriers which are reduced to facilitate nucleation processes. While mechanisms and theories have been postulated for the formation of surface precipitates, many researchers have reported indirect (macroscopic data) and direct (spectroscopic data) evidence for aluminum‐ and iron‐P surface precipitates on metal oxides and clays. Chen et al. (1973) indirectly suggested a nucleation of a new P‐containing phase (i.e., P surface precipitate) at the kaolinite–water interface. They observed a disequilibrium phenomenon for the P adsorption envelope on kaolinite where the adsorption maximum at pH 4 increased with increasing time from 1 day to 23 days when P in the bulk solution was undersaturated with respect to aluminum phosphate. van Riemsdijk and Lyklema (1980) also suggested that P adsorption on gibbsite resulted in the formation of a potassium aluminum phosphate precipitate. This was due to: (1) the increase in the rate of excess sorption (i.e., beyond monolayer coverage) with increasing supersaturation, (2) enhanced P adsorption with increasing potassium in the bulk solution, and (3) increased surface area after P adsorption (van Riemsdijk and Lyklema, 1980). Ler and Stanforth (2003) reported on changes in the Zeta (x)‐potential of P sorbed on goethite at pH 5

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with time. The x values initially dropped due to chemisorption of P, but gradually returned with time to those of goethite without sorbed P. They suggested that the long‐term reactions most likely involved the dissolution of ferric ions from goethite surfaces and the subsequent reaction between ferric (aq) and surface‐bound P, resulting in the formation of ferric phosphate surface precipitates. Direct evidence for P surface precipitates has been suggested from microscopic and spectroscopic investigations. Electron microscopy, electron diVraction, and XRD were utilized to investigate long‐term (40 days) P sorption on amorphous Al(OH)3 and a‐Al2O3 at pH 5. Researchers observed the formation of a new solid aluminum‐phosphate mineral (Sterrettite‐like) at aluminum oxide surfaces (van Riemsdijk et al., 1977). Nanzyo (1984, 1986) investigated P adsorption on amorphous aluminum and iron oxides over a wide pH range (4–9) using diVuse reflectance infrared spectroscopy. He suggested that the formation of amorphous aluminum phosphate‐like and amorphous iron phosphate‐like products increased with decreasing pH based on the observation of similar IR spectra between the adsorption complex and synthetic minerals (i.e., aluminum phosphate and iron phosphate). XPS, STEM, and XRD investigations have shown that a mixture of goethite and tinticite (Fe4(PO4)4(OH)6 7H2O) are formed after 18 days of P adsorption (25 C and [P]i ¼ 0.01 mM liter1) at pH 3 (Jonasson et al., 1988). Multispectroscopic techniques (Auger, XPS, scanning SIMS, and TEM) have been applied to investigate P adsorption on natural goethite containing SiO2, Al2O3, and Mn2O3 impurities ([P]i ¼ 1 mM, 90 days, and 60 C) (Martine et al., 1988). The mineral griphite (Fe3Mn2(PO4)2.5(OH)2) was found in isolated crystallites. Martine et al. suggested that P fixation in soils may be controlled by precipitation reactions on Fe oxides. Similarly, Wang and Tzou (1995) observed an Fe‐P precipitate‐like compound on a P‐reacted hematite surface. Ex situ Mo¨ssbauer analysis showed quadrupole splitting for P‐adsorbed hematite ([P]i ¼ 155 mg g1, pH 3, 24 h, and 298 K). Such splitting resembled the spectra observed for vivianite (Fe(II)3(PO4)2) and strengite (Fe(III)(PO4)) (McCammon and Burn, 1980; Wang, 1987). Since XRD analysis did not detect the mineral phase, the mineral phase was presumed to be amorphous. Slow P adsorption on allophane with continuous silicon release has been indirectly linked to a precipitation reaction by many researchers. A reaction with P caused a disruption in the allophane structure by displacing structural silicon. Exposed reactive sites might facilitate P precipitation as aluminum phosphate (Nanzyo, 1987; Parfitt, 1989; van Riemsdijk and Lyklema, 1980). Induced silicon release during phosphate adsorption has also been reported on soils and natural clay minerals (i.e., kaolinite, natural allophane, ferrihydrite, goethite, and allophanic clay) (Low and Black, 1950; Mattson and Hester, 1935; Parfitt, 1989; Reifenberg and Buckwold, 1954). Rajan (1975) observed silicon release during P adsorption on allophanic clay which was

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pretreated with selenite to remove labile silicate. P isomorphic substitution of structural silicon in allophanic clay has been also suggested. Veith and Sposito (1977) also reported the formation of aluminum phosphate surface precipitates on allophatic materials after P adsorption. The formation of X‐ray amorphous analogues of variscite and Na‐montebrasite were demonstrated by reacting Al2O3 nH2O, synthetic allophanes and allophanic soils with P at varying acidity. The formation of amorphous Al‐phosphates was favored by both the large values of the hydration number in Al2O3 nH2O and the increase in the acidity of the added P solution. The formation of amorphous Al‐phosphate by secondary precipitates has been indicated by: (1) an observed slow reaction rate between P and Al coatings, (2) the immediate and significant increase of silicate in solution when aluminosilicate minerals were reacted, and (3) a significant amount (above monolayer coverage) of P adsorption. While numerous macroscopic, microscopic, and spectroscopic studies have speculated on the formation of P surface precipitates at mineral surfaces, there are some questions about these findings. These doubts have been raised because: (1) macroscopic data do not directly suggest a mechanism for surface precipitate formation, (2) high P concentration (>1 mM) might cause a supersaturated condition with respect to bulk precipitates, and (3) ex situ conditions in XRD, XPS, SEM, and TEM analyses might create artifacts with residual P in the samples. Macroscopic and microscopic data must be carefully combined with in situ spectroscopic techniques to draw better conclusions about the formation of surface precipitates.

F.

HIGHER ENERGY BINDING THROUGH CHEMICAL RECONFIGURATION

Real soil environments are not in equilibrium and always undergo slow chemical changes to reach equilibrium (Koskinen and Harper, 1990; Steinfield et al., 1989). As the materials become more stable (i.e., lowering solubility or resulting multidentate complexes) with increasing aging time through transformations, the release of the adsorbate becomes slower, as evidenced by slow desorption phenomena. Lowering the Gibbs free energy with increasing entropy converts the products to more stable compounds. As a result, the activation energy of the desorption phenomena can be expressed as Ed ¼ Ea þ Q. This theory can be categorized into two groups: (1) a chemical transformation from two‐dimensional surface complexes to three‐dimensional precipitates and (2) a reconfiguration within two‐dimensional surface complexes. Stumm and Morgan (1995a) suggested a chemical transformation theory, the Oswald‐step rule. This theory proposes that the nucleation process involves the formation of the least stable solid phase (highly soluble amorphous materials) first followed by the formation of a more stable solid phase (insoluble crystalline materials).

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The stability of the surface complexes can be altered from monodentate to multidentate configuration with increasing time. The more the bonds that are associated with the surface functional group of the solid, the greater the stability. The chemical transformation from mononuclear to binuclear bridging complexation was suggested to describe the hysteresis phenomena for P adsorption on soils and soil components (Kafkafi et al., 1967; Mengel, 1985); however, there is no molecular scale evidence to support this chemical reconfiguration theory.

VI.

FUTURE RESEARCH NEEDS

This chapter has described past and recent research eVorts to understand P retention and release mechanisms in soils and soil minerals. Although numerous P adsorption and desorption studies and several benchmark spectroscopic studies have provided vital information about P reactivity in soils and sediments, uncertainties still persist in predicting the long‐term fate and transport of P in soil–water environments. An historical understanding of P reactivity on model adsorbents should lead to new investigations of: (1) P reaction dynamics in more complex model systems (e.g., mixed adsorbents and adsorbates) and (2) P solid‐state speciation in high‐P‐containing soils and sediments. Although several researchers have begun to investigate the eVects of competitive ligands on P reactivity (Borggaard et al., 2005; Geelhoed et al., 1997, 1998; Hawke et al., 1989; Hongshao and Stanforth, 2001; Johnson and Loeppert, 2006; Liu et al., 1999), our understanding of P reactivity with respect to P speciation in natural materials is still limited to operationally defined extractable fractions via chemical extractions. There are not many studies characterizing P solid‐state speciation in natural materials. Using modern microscopic and spectroscopic techniques (e.g., STEM, synchrotron‐based X‐ray fluorescence spectroscopy and synchrotron‐based XRD), P solid‐state speciation in soils and sediments can be better characterized, and these research findings will lead to a better understanding of the P retention and release in heterogeneous soils and mixed‐model adsorbent and ligand systems. Such comprehensive research results could be helpful in designing more eVective in situ remediation technologies and nutrient management programs to enhance environmental quality.

REFERENCES Aharoni, C., and Suzin, Y. (1982). Application of the Elovich equation to the kinetics of chemisorption. Part 3. Heterogeneous microporosity. J. Chem. Soc., Faraday Trans. 1 78, 2329–2336.

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Ryden, J. C., and Syers, J. K. (1977). Desorption and isotopic exchange relationship of phosphate sorbed by soils and hydrous ferric oxide. J. Soil Sci. 28, 596–609. Ryden, J. C., Syers, J. K., and Harris, R. F. (1973). Phosphorus in runoV and streams. In ‘‘Advances in Agronomy’’ (N. C. Brady, Ed.), Vol. 25, pp. 1–45. Academic Press, New York. Ryden, J. C., Mclaughlin, J. R., and Syers, J. K. (1977a). Time‐dependent sorption of phosphate by soils and hydrous ferric oxides. J. Soil Sci. 28, 585–595. Ryden, J. C., Syers, J. K., and Mclaughlin, J. R. (1977b). EVects of ionic strength on chemisorption and potential‐determining sorption of phosphate. J. Soil Sci. 28, 62–71. Sanchez, P. A., and Uehara, G. (1980). Management considerations for acid soils with high phosphorus fixation capacity. In ‘‘The role of phosphorus in agriculture’’ (F. E. Khasawneh, E. C. Sample, and E. J. Kamprath, Eds.), American Society of Agronomy, Madison, Wisconsin. Sato, S., Solomon, D., Hyland, C., Ketterings, Q. M., and Lehmann, J. (2005). Phosphorus speciation in manure and manure‐amended soils using XANES spectroscopy. Environ. Sci. Technol. 39, 7485–7491. Sei, J., Jumas, J. C., Olivier‐Fourcade, J., Quiquampoix, H., and Staunton, S. (2002). Role of iron oxides in the phosphate adsorption properties of kaolinites from the Ivory Coast. Clays Clay Miner. 50, 217–222. Shang, C., Stewart, J. W. B., and Huang, P. M. (1992). pH eVect on kinetics of adsorption of organic inorganic phosphates by short‐range ordered aluminum and iron precipitates. Geoderma 53, 1–14. Shapiro, R. E. (1957). The eVect of flooding on the availability of P and nitrogen. Soil Sci. 85, 190–197. Sharpley, A. N., and Rekolainen, S. (1997). Phosphorus in agriculture and its environmental implications. In ‘‘Phosphorus Loss from Soil to Water’’ (H. Tunney, O. T. Carton, P. C. Brookes, and A. E. Johnston, Eds.), pp. 1–54. Cab International, New York. Sharpley, A. N., Foy, B., and Whiters, P. (2000). Practical and innovative measures for the control of agricultural phosphorus losses to water: An overview. J. Environ. Qual. 29, 1–9. Sheppard, S. C., and Racz, G. J. (1984). EVects of soil temperature on phosphorus extractability. 1. Extractions and plant uptake of soil and fertilizer phosphorus. Can. J. Soil Sci. 64, 241–254. Snoeyink, V. L., and Jenkins, D. (1980). Precipitation and dissolution. In ‘‘Water Chemistry,’’ pp. 243–315. John Wiley & Sons, Inc., New York. Sparks, D. L. (1987). Dynamics of soil potassium. In ‘‘Advances in Soil Science’’ (B. A. Stewart, Ed.), Vol. 6, pp. 1–63. Springer Verlag, New York. Sparks, D. L. (1989). ‘‘Kinetics of Soil Chemical Processes.’’ Academic Press, Inc., San Diego. Sparks, D. L. (1995a). ‘‘Environmental Soil Chemistry.’’ Academic Press, Inc., San Diego. Sparks, D. L. (1995b). Sorption phenomena on soils. In ‘‘Environmental Soil Chemistry,’’ pp. 99–139. Academic Press, San Diego, CA. Sposito, G. (1986). Distinguishing adsorption from surface precipitation. In ‘‘Geochemical Processes at Mineral Surfaces’’ (J. A. Davis and K. F. Hayes, Eds.), pp. 217–228. American Chemical Society, Washington, DC. Sposito, G. (1989). Soil adsorption phenomena. In ‘‘The Chemistry of Soils.’’ Oxford University Press, New York, Oxford. Steinfield, J. I., Francisco, J. S., and Hase, W. L. (1989). ‘‘Chemical Kinetics and Dynamics.’’ Englewood CliVs, Prentice Hall. Stumm, W., and Morgan, J. J. (1995a). ‘‘Aquatic Chemistry, Chemical Equilibria and Rates in Natural Waters.’’ John Wiley & Sons, Inc., New York. Stumm, W., and Morgan, J. J. (1995b). Theory of elementary processes. In ‘‘Aquatic Chemistry, Chemical Equilibria and Rates in Natural Water,’’ pp. 69–76. John Wiley & Sons, Inc., New York.

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Tejedor‐Tejedor, M. I., and Anderson, M. A. (1990). Protonation of phosphate on the surface of goethite as studied by CIR‐FTIR and electrophoretic mobility. Langmuir 6, 602–611. Toor, G. S., Hunger, S., Peak, J. D., Sims, J. T., and Sparks, D. L. (2006). Advances in the characterization of phosphorus in organic wastes: Environmental and agronomic applications. In ‘‘Advances in Agronomy’’ (D. L. Sparks, Ed.), Vol. 89, pp. 1–72. Elsevier, San Diego. Torrent, J. (1991). Activation energy of the slow reaction between phosphate and goethites of diVerent morphology. Aust. J. Soil Res. 29, 69–74. Torrent, J., Barron, V., and Schwertmann, U. (1990). Phosphate adsorption and desorption by goethite diVering in crystal morphology. Soil Sci. Soc. Am. J. 54, 1007–1012. Torrent, J., Schwertmann, U., and Barron, V. (1992). Fast and slow phosphate sorption by goethite‐rich natural materials. Clays Clay Miner. 40, 14–21. Vaithiyanathan, P., and Correll, D. L. (1992). The Rode river watershed: Phosphorus distribution and export in forest and agricultural soils. J. Environ. Qual. 21, 280–288. van der Zee, S. E. A. T. M., and van Riemsdijk, W. H. (1988). Model for long‐term phosphate reaction kinetics in soil. J. Environ. Qual. 17, 35–41. van Riemsdijk, W. H., and de Haan, F. A. M. (1981). Reaction of orthophosphate with a sandy soil at constant supersaturation. Soil Sci. Soc. Am. J. 45, 261–266. van Riemsdijk, W. H., and Lyklema, J. (1980). Reaction of phosphate with gibbsite (Al(OH)3) beyond the adsorption maximum. J. Colloid Interface Sci. 76, 55–66. van Riemsdijk, W. H., Weststrate, F. A., and Beek, J. (1977). Phosphate in soils treated with sewage water: III. Kinetics studies on the reaction of phosphate with aluminum compounds. J. Environ. Qual. 6, 26–29. van Riemsdijk, W. H., Boumans, L. J. M., and de Haan, F. A. M. (1984). Phosphate sorption by soils: I. A model for phosphate reaction with metal‐oxides in soil. Soil Sci. Soc. Am. J. 48, 537–541. Veith, J. A., and Sposito, G. (1977). Reactions of aluminosilicates, aluminum hydrous oxides, and aluminum oxide with o‐phosphate: The formation of x‐ray amorphous of variscite and montebrasite. Soil Sci. Soc. Am. J. 41, 870–876. Wang, M. K. (1987). Synthetic and characterization of iron phosphate and aluminum‐ substituted iron phosphate compounds. J. Chin. Agric. Chem. Soc. 25, 398–411. Wang, M. K., and Tzou, Y. M. (1995). Phosphate sorption by calcite, and iron‐rich calcareous soils. Geoderma 65, 249–261. Wang, H. D., Harris, W. G., and Yuan, T. L. (1991). Noncrystalline phosphates in Florida phosphatic soils. Soil Sci. Soc. Am. J. 55, 665–669. Waychunas, G. A., Rea, B. A., Fuller, C. C., and Davis, J. A. (1993). Surface chemistry of ferrihydrite: Part 1. EXAFS studies of the geometry of coprecipitated and adsorbed arsenate. Geochim. Cosmochim. Acta 57, 2251–2264. Willett, I. R., Chartres, C. J., and Nguyen, T. T. (1988). Migration of phosphate into aggregated particles of ferrihydrite. J. Soil Sci. 39, 275–282. Withers, P. J. A. (1996). Phosphorus cycling in UK agriculture and implications for water quality. Soil Use Manage. 12, 221–228. Wong, J., Lytle, F. W., Messmer, R. P., and Maylotte, D. H. (1984). K‐edge absorption spectra of selected vanadium compounds. Phys. Rev. B 30, 5596–5610. Yao, W., and Millero, F. J. (1996). Adsorption of phosphate on manganese dioxide in sea water. Environ. Sci. Technol. 30, 536–541.

ECOLOGICAL AGRICULTURE IN CHINA: PRINCIPLES AND APPLICATIONS Huixiao Wang,1 Longhua Qin,1 Linlin Huang1 and Lu Zhang2 1

Key Laboratory for Water and Sediment Sciences, Ministry of Education, College of Water Sciences, Beijing Normal University, Beijing 100875, People’s Republic of China 2 CSIRO Land and Water, Canberra, ACT 2601, Australia

I. Introduction A. Alternative Agriculture B. Ecological Agriculture in the West C. International Sustainable Agriculture II. Chinese Ecological Agriculture A. Environmental Problems in China B. Researches on SA in China C. Characteristics of CEA D. Basic Principles of CEA III. Development and Achievements of CEA IV. Practical Aspects of CEA A. Vertically Distributed Farming B. Multilevel Organic Substance Utilization C. Energy Exploitation D. Integrated Control Techniques E. Introducing New Varieties F. Comprehensive Management of the Agricultural Environment V. Case Studies of CEA A. Ecological Agriculture in Mountainous Regions B. Water‐Collecting Ecological Agriculture in Western China C. From Ecological Agriculture to Ecological Industry VI. Problems A. From Theory to Practice B. Poverty C. Lack of Systematic Theoretical Research D. Small Production Scale E. Lack of Funds F. Education of Farmers G. Lack of Market Competiveness VII. Concluding Remarks A. Further Studies on Theories of CEA B. Modern Techniques Application C. Organic Food: Filling a Gap in the Market D. Ecological Agriculture and Township Enterprise 181 Advances in Agronomy, Volume 94 Copyright 2007, Elsevier Inc. All rights reserved. 0065-2113/07 $35.00 DOI: 10.1016/S0065-2113(06)94004-8

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China, as a developing country, faces a unique combination of overpopulation, shortage of cultivated land, and a stretched economy. Compounding the problem, industrialization and overuse of natural resources has resulted in serious pressures on agriculture and environment. In order to solve issues regarding the environment, resources, energy, and food, agricultural development needs to be approached in a new way, and this chapter identifies a uniquely Chinese ecological agriculture (CEA), a sustainable development strategy that has significant potential as a way forward. This chapter takes an overview of ecological agriculture in China and the current status of research in the field. The key characteristics of CEA are discussed, and similarities and diVerences are set out with the usual alternative agriculture, ecological agriculture as known in the West, and with international sustainable agriculture. This chapter is based on studies extending over more than 20 years following the first introduction of CEA, and the accumulated principles and practices of CEA are summarized. Three case studies, representing typical ecological agriculture patterns, are discussed: ecological agriculture in mountainous regions, water‐collecting ecological agriculture in western China, and eVorts to move from ecological agriculture to ecological industry. Finally, problems hindering the development of CEA, and avenues for future research, # 2007, Elsevier Inc. are pointed out.

I. INTRODUCTION The world currently faces serious pressures from population growth, food shortages, resource depletion, energy demands, and pollution. In this context, governments must seek a balance between agriculture development, consumption of natural resources, and environmental protection. After the 1940s, modern petrochemical‐based agriculture—an intensive agricultural system—accelerated agricultural development in developed countries. A feature of petrochemical‐based agriculture is the widespread use of industrial products such as oil, chemical fertilizer, pesticides, machinery, electricity, and so on. Some of the negative eVects have included excessive consumption of energy, degradation of agricultural ecosystems, compromised food safety, and increasing environmental pollution. In the mid‐1970s, therefore, many types of alternative agriculture were advocated in order to avoid these drawbacks.

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A. ALTERNATIVE AGRICULTURE Alternative agriculture comes in many forms, but includes ecological agriculture, biological agriculture, organic agriculture, conservational farming, biodynamic agriculture, regenerative agriculture, natural agriculture, and low‐energy agriculture, among others (Francis et al., 1986; Hodges, 1982; Lockeretz, 1988; Lockeretz et al., 1981; Merrill, 1983). Organic farming emphasizes recycling of farm‐generated nutrient source, limits bringing in nutrients in the form of livestock feeds, and avoids chemical fertilizers (Scofield, 1986). A study by Lockeretz et al. (1981) on 174 organic farmers in the American Corn Belt showed that, compared to conventional methods, organic methods consumed less fossil energy and caused less soil erosion, but had mixed eVects on soil nutrient status and grain protein content. In general, alternative agriculture tries to establish a self‐contained agricultural ecosystem. It aims for system balance and protection of the environment in the process maintaining agricultural production and quality. It rejects the use of synthetic chemicals, passing over the increased production and profits they oVer. It tolerates low agricultural eYciency and diYculty in large‐scale implementation in favor of long‐term sustainability. Wang et al. (2004) consider that alternative agriculture is an extreme response to modern petrochemical agriculture.

B.

ECOLOGICAL AGRICULTURE IN

THE

WEST

The term ‘‘ecological agriculture’’ was first put forwarded by W. Albreche, an American pedologist, in 1970. Ecological agriculture was defined as a small‐scale agricultural system displaying environmental, ethical, and aesthetic aspects. In 1981, the British agronomist M. Worthington described characteristics of ecological self‐suYciency, low input, and high economic vitality. A type of alternative agriculture, Western ecological agriculture carries the central idea of trying to operate agriculture on the basis of ecology, instead of chemistry. ‘‘Ecological’’ refers to the principles and processes that govern the natural environment (Lockeretz, 1988). Ecological agriculture in the West has been established against a background of agricultural overproduction and serious environmental pollution; some of its adherents believe that soil and water pollution in fact threaten human health and survival. Ecological agriculture is based on the principle of avoiding pollutant sources in the food chain and preventing soil and water pollution. The production system is relatively simple, usually based on only growing crops or sometimes a system of crops and barn‐fed animals. It selectively uses chemical fertilizers, but strictly rejects pesticides, herbicides, hormones, and the like. Because it emphasizes the environmental benefits at the expense of reduced gross output and

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profits, the price of the resulting ecological products is much higher than that of conventional agricultural products (Shen et al., 1993). These characteristics have restricted the implementation of ecological agriculture in the West. After 20 years of use, the area under ecological agriculture in Germany is now only 3.2% of the nation’s farmland (Cheng, 2002). Perhaps one way out of the dilemma is to diversify changing ecological agriculture from a single ecological structure to one with multiple components. This idea, which China is moving toward, has the potential, we believe, to lead to a renaissance of ecological agriculture in the West.

C. INTERNATIONAL SUSTAINABLE AGRICULTURE In view of the production limitations of alternative agriculture, sustainable agriculture (SA) has sought to harmonize the triple benefits of environment, economy, and society. It was put forward in the mid‐1980s and called the second replacement of modern petrochemical agriculture (Liu, 1995). SA, an integrated production pattern of agriculture, diVers from ecological agriculture, connoting much more than ecological agriculture (Chai et al., 2005). SA arose from the drawbacks of alternative agriculture. Although great achievements had been made in many aspects of alternative agriculture— saving energy, protecting natural resources, improving the environment, and providing pollution‐free food—its low material inputs generally meant low yields, low economic benefits, and low labor productivity, features that, as we have said above, go against the high eYciency required by modern society. Jin and Jin (1999) describe how alternative agriculture cannot be extended to a large area and how this low‐level sustainable agricultural production pattern cannot meet the demands of economic development— even in the developed countries, the scheme only accounts for 0.3% of total cultivated land in the developed countries. SA was developed through applying sustainable development ideas to agriculture production. Brown (1981), an American agricultural scientist, systematically set out the principles of sustainable development from which a theoretical basis was established. Douglass (1984) also advocated agricultural sustainability. Francis et al. (1988) defined a sustainable agricultural system as a management strategy that helps the producer reduce costs of inputs, minimizes the impacts of the system on the immediate and oV‐farm environment, and provides a sustained level of production and profit from farming. Later, in America, low‐input sustainable agriculture (LISA) and high‐eYciency sustainable agriculture (HESA) were formulated (Lee and Baum, 1999; Walter and Young, 1987). In April 1991, FAO organized an international conference on agriculture and the environment in the Netherlands, from which the concept of ‘‘sustainable agriculture and rural development’’ (SARD) arose.

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II. CHINESE ECOLOGICAL AGRICULTURE In China, much work has been done on improving the ecology of the agricultural environment, a remarkable achievement in a country that has only about 7% of the world’s cultivated land but supports 22% of its population (Guo, 1997). As a developing country, China faces a unique combination of overpopulation, shortage of cultivated land, and a stretched economy. Chai et al. (2005) point out that, per capita, the country’s cultivated land, forest, grassland, and water resources are only a small fraction of the world average (1/3, 1/5, 1/4, and 1/4, respectively). Compounding the problem, industrialization and overuse of natural resources has resulted in serious pressures on agriculture and environment. Population increases have created a serious burden to economic and social development in China, especially for agriculture, and particularly in the poor and mountainous regions. As a result, agriculture in China is in a state lacking reserves for further development.

A. ENVIRONMENTAL PROBLEMS

IN

CHINA

Deterioration in the agricultural environment is shown by the following considerations. (1) Soil erosion is serious, occupying an area of 38% of the country, and this results in large nutrient losses from arable land every year. (2) Land subject to desertification has reached 33 million ha, and this increases by 0.234 million ha each year. (3) The combined area aVected by desertification, salinization, and grassland degradation has reached 87 million ha, with the result that animal husbandry has low productivity. Meat products originating from pastured regions only accounts for 5% of the country’s total. (4) Farmland polluted by waste gas, waste water, and residues from large‐scale industrial and mining enterprises, together with that from rural industries, totals 10 million ha. The area aVected by acid rain extends each year, producing serious damage to soil and water resources. (5) During the country’s industrial development, large areas of fertile land were occupied and taken out of production. This factor contributed to the cultivated land per person decreasing from 0.16 hm2 in the 1950s to 0.08 hm2 at present. (6) Excessive groundwater extraction for industrial and agriculture purposes has resulted in lowering of groundwater tables, land subsidence, and secondary salinization. The area of the salinized land has reached 27 million ha, aVecting 7 million ha of cultivated land. (7) The liberal use of chemical fertilizers has lead to reduced soil organic matter and increased nitrate concentration of groundwater. (8) Large‐scale use of agricultural pesticides has increased insect and pesticide resistance, and contributed to residues in agricultural products as well. (9) Due to the development of

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large‐scale animal husbandry in rural areas surrounding big cities, serious groundwater pollution has resulted from the associated large volumes of eZuent. Given this situation, development for the sake of pursuing fast economic growth at the expense of the environment cannot continue. All the above‐ mentioned problems reflect a Chinese agricultural eco‐crisis, the long‐term harm of which is more serious than any short‐term financial deficit. To overcome the problems, the biological solutions we seek must be inexpensive, safe, and sustainable. Against this background we seek to establish and develop Chinese ecological agriculture (CEA). CEA employs recycling to develop an economy based on material recycling, a pattern required by any sustainable development strategy. Recycling means reducing the amount of resources used and the quantity of waste discharged. For eYcient recycling, all waste must be reused. When applied to ecological agriculture—mixed farming, crop cultivation, stockbreeding, agricultural products processing, tourism, and so on—every waste stream will have to be captured and recycled.

B. RESEARCHES

ON

SA IN CHINA

Studies of the theory and practice of SA in China have been continuing for more than 10 years (Cheng, 2002; Liu, 1995, 2004; Liu and Wu, 2000, 2001; Liu and You, 1993; Sun and Hu, 1993; Wang, 1993). Wang (1993) defined SA as an overall agricultural production system balancing yields, benefits, crop quality, and the environment. It is a type of SA which is able to supply agricultural products for the present generation without damaging resources or the environment for future generations. The characteristics of SA are production sustainability, ecological sustainability, economic sustainability, social sustainability, intensive use of agricultural technology, and multiple objectives (Shen et al., 2002). Relevant theories for SA include agricultural systems theory, ecological economic theory, and system cybernetics (Liu and Wu, 2000). Internationally, SA was first appeared in the developed countries after they struck problems in conventional agriculture. Because agricultural production levels in these regions are commonly high, the major objective of SA is to protect resources and the environment; however, for developing countries like China, the idea of appreciably reducing yields is not appropriate. For developing countries, growth is a priority, and at the same time more attention must be paid to rational use of natural resources and environmental protection (Lu, 1995); in this way, the goal is that agriculture can be expanded and made sustainable. To develop SA, Liu and You (1993) considered that two aspects need to be considered: first, the sustainability of any agricultural resources exploitation and use, and second, the adaptability of agricultural production to social

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Ecological environment Agricultural natural resources

Rational exploitation and utilization Management mechanism and policy system

Input of science and technology

Rational input of material and energy

Sustainable agriculture

High yield, top quality, and high efficiency

Economic, environmental, social targets Concordant relation of supply and demand Perfection of ecology and environment

Social demands for agriculture

Population, economic development

Figure 1 Interactions in sustainable agriculture.

demands. The former are those natural agricultural resources such as water, land, climate, and organisms; the latter includes social economy, ecology, and the environment. The relationships are shown in Fig. 1. Wang (1993) considered that abundant natural resources, rational input of material and energy, economic development, input of science and technology, management, and policy form the basis for SA—materially, economically, technologically, and socially. Factors that could contribute to making Chinese agriculture unsustainable include the following: overpopulation, decreasing area and quality of cultivated land, soil erosion, desertification, water resources problems, environmental pollution, shortage of farming capital, lack of science and technology inputs, diYculty in employing surplus rural labor, ineYcient production systems and industrial structures, low agricultural profits compared with other industries, and outmoded management systems (Liu, 2004; Shen et al., 2002).

C. CHARACTERISTICS

OF

CEA

Owing to its limited economy and huge area, China cannot follow the way of ‘‘industrial agriculture’’ or ‘‘oil (petrochemical) agriculture.’’ Moreover, ‘‘alternative agriculture’’ will not work in China, where agricultural production must be kept high. During the transition from traditional agriculture to modern agriculture, a sustainable development pattern with uniquely Chinese

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characteristic—CEA—has gradually taken shape. As foreshadowed, ecological agriculture in the West emphasizes low inputs or absolutely rejects the use of agricultural chemical products. CEA, a derivative of SA, is diVerent: it emphasizes high land productivity, pursuing high primary production (plant photosynthesis) and putting measured amounts of energy into the agricultural ecosystem. Consequently, some inputs of material and energy are needed to increase eYciency; at the same time it seeks to reduce the cost and dependence of agricultural ecosystem on other inputs, and to heavily reduce the likelihood of soil and water pollution from residues of chemical fertilizers, pesticides, and excrement. Ecological agriculture in the West is designed for production on a family or a small‐farm scale, whereas the emphasis of CEA is on large‐scale farming, forestry, animal husbandry, sideline production, and fisheries. It uses ecological principles and aims for a rational layout and design of the whole agricultural system, sometimes even on a regional scale (Cheng, 2002). The idea of CEA came about in the early 1980s. On the basis of using the ideas of foreign alternative agriculture for reference, it has a long background in traditional organic agriculture, to which it owes its primary concepts and evolving processes. The aim of CEA is to coordinate the relationship between man and nature, to raise agricultural production, and support rural economic and socially sustainable development. The specific objectives of CEA are optimizing production, avoiding pollution, having a high eYciency of resource utilization, producing healthy and residue‐free agricultural products, and protecting rural environment. It achieves this through a combination of industrial diversity, pollution control, recycling, increasing input eYciency, organic and green production methods, and sympathetic use of the rural environment (Zhang and Gao, 2004). Ma and Li (1987) defined ecological agriculture as the application of ecological engineering to agriculture, giving an agricultural production system that uses the ecological principles of organism mutualism and material recycling. It combines systems engineering methods with modern science and technology, rationally interlinking a mixture of farming, forestry, animal husbandry, sideline production, and fisheries according to the natural resources locally available. The outcome is a symbiosis, a production of economic, ecological, and social benefits larger than could be obtained from its individual components.

D.

BASIC PRINCIPLES

OF

CEA

The agricultural ecosystem is a complex biological system that includes the human being, who is both the consumer and manager of the ecosystem. Ecological agriculture aims to coordinate the system’s components to achieve

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the largest system stability, the least artificial input, and the most comprehensive benefits from the viewpoint of ecology, economy, and society. The theoretical basis of ecological agriculture lies in the principles of ecology and ecological economics. One of the primary founders of CEA summarized its basic concept (Ma, 1983) as ‘‘unity, harmony, circulation, recycling,’’ and the specific principles are set out below (Jin and Jin, 1999; Li et al., 2001). 1.

Principle of Adaptation Between Organisms and Environment

The two components of an ecosystem, organisms and environment, bear a close relationship and undergo complex material and energy exchanges. First of all, the environment provides the essential materials and energy for organisms to live and reproduce—air, light, water, heat, nutrition, and so on. At the same time, during the processes of living, reproduction, and movement, organisms constantly return material to their environment through diVusion, excretion, or other means. So organisms and environment continuously interact and adapt. According to this principle, they adapt over time to local conditions provided there is a sympathetic balance between species and varieties. Therefore, a suitable combination of land cultivation, rotations, and soil improvement should augment agricultural production. Pushed too far, however, environmental quality may be degraded. 2. Principle of Interdependence Between Organisms Organisms in an ecosystem are interdependent and, through nutrition relationships, exert mutual control so as to form a complex food net of green plants, herbivorous animals, and carnivorous animals. Changes in one of the chains will influence the others or even the total food net. There is a strict quantitative relationship between the organisms in the food chain: organisms at adjacent positions in the food chain bear a definite ratio in number of individuals, biomass, or energy. Following this principle, in CEA the organisms in a food chain should be connected so as to exploit resource potential, and hence agricultural production, to the greatest degree. Of course, just connecting components at random is likely to destroy the ecological balance. 3.

Principle of Multilevel Energy Utilization and Material Recycling

An ecosystem’s food chain predetermines how energy flows and materials transform. Ecosystems fail when, through overexploitation of some renewable resource, overproduction of wastes occurs. This disturbs and slows down the normal nutrient cycling, resulting in an unbalance of materials,

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190 Solar radiation

Grain

Fertilizer

Crop

Manure Farm animals Sacchariferous feed Straw

Inoculation Scraps of Edible fungus bed Inoculation fungus

Products Earth worm Products

Frass and scraps

Products

Figure 2 Multilevel utilization of crop straw. Adapted from Ma (1983).

and finally pollution and environmental disruption (Cao et al., 2002). Reclaiming wastes enhances opportunities for material production and organism reproduction. For example, returning straw back to a field is an eVective way of retaining soil organic matter (Ma, 1983). However, if straw is returned to the soil directly, its fertilizer eYciency can only be brought into play after a long period of fermentation and decomposition. But bringing ecological principles to bear, the straw can be inserted into a food chain, increasing soil fertility and reducing pollution. The key is to feed the straw to cattle: after a process of saccharification or ammonification, crop straw can become cattle food, the excreta of domestic animals can then be used to cultivate edible fungus, the residual fungus bed can be used to feed earthworms, and finally the earthworm residues can be returned to the field as fertilizer. Material is thus used at many levels, greatly improving energy eYciency (Fig. 2).

4.

Principle of Structural Stability and Harmonization of Function

In natural ecosystems, a relatively stable structure among organisms is built via long‐term interaction between organisms and the environment. Ecological agriculture aims to grow products with top quality and high yield, and to create fertile land and an attractive living environment.

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A stable agricultural ecosystem must ensure that the system continues to run normally. The biological components in the agricultural ecosystem are elaborately arranged according to the production target and ecological principles such as symbiosis among organisms. Good examples of ecological agriculture include the combination of fruit tree cultivation and beekeeping, fish culture in paddy fields, using rhizobium of leguminous plants to fix nitrogen, and improving soil structure.

5.

Principle of Integration of Ecological and Economic Benefits

Ecological agriculture, like other types of agriculture, is a human economic activity, the aim of which is to increase output and economic revenue. In order to simultaneously achieve high economic and ecological benefits in ecological agriculture, the above‐mentioned principles must be used: rational arrangement of resources, full use of labor resources, rationalization of economic structure, and specialization and socialization of ecological agriculture (Li et al., 2001).

III.

DEVELOPMENT AND ACHIEVEMENTS OF CEA

In the 1970s, it was proposed that the concept of ecological balance should be used to guide agricultural research. The term ‘‘ecological agriculture’’ was used in China for the first time in 1980 at a symposium of national agricultural ecological economics held in Yinchuan, Ningxia. The symposium considered ecological agriculture to be a new type of integrated agricultural system harmoniously blending agricultural production, rural economic development, environmental protection, and high‐eYciency utilization of natural resources. In the early 1980s, the environmental protection group of the State Council launched a pilot project on ecological agriculture. From that beginning, more than 2000 counties, towns, and villages in the whole country have now constructed ecological agriculture systems. The systems are individually tailored so that each technological pattern fits diVerent natural ecological conditions and social economic development levels (Bian, 2000). Development was more rapid after the 1990s when the Ministry of Agriculture, the State Planning Commission, the State Science and Technology Commission, the Ministry of Finance, the Ministry of Water Resources, the Ministry of Forestry, and the State Bureau of Environmental Protection jointly launched a pilot project of national ecological agriculture at the county level. According to the statistics from the first 51 state‐level ecological agriculture experimental counties, the annual growth rates of gross domestic product, gross agriculturaloutput value, and the net income per worker rose by 2.2%, 0.6%, and

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1.5% (respectively) for these counties compared to the national average over the same period. Control rates of soil erosion and desertification were 73.4% and 60.5%, respectively, and tree coverage raised by 3.7%. Some seven ecological agriculture experimental units have already been awarded inclusion in ‘‘500 excellent works of the global environment protection’’ conferred by the United Nations Environment Program (UNEP). On the basis of the achievements of the first 51 state‐level ecological agricultural pilot counties, the second and third groups of 100 demonstration counties were organized and ecological agriculture implemented at 500 provincial‐level counties (Li, 2004). The Chinese central government has paid close attention to ecological agriculture, and it has been propagated and popularized extensively; it has tended to be combined with rural reform and the structural readjustment of agricultural industries. In ‘‘Decision on environmental protection work of the State Council’’ (1984), it was stated that ‘‘it is needed to protect the ecological environment earnestly, to extend ecological agriculture actively, and to prevent the agricultural environment from pollution and destruction.’’ In ‘‘Ten‐year planning for national economy and social development and the eighth five‐year plan (1991),’’ it was clearly pointed out that environmental demonstration engineering and ecological agriculture construction should be promoted. In 1992, the State Council regarded ecological agriculture as 1 of 10 major countermeasures for environment and development in China and called for an increase in construction of ecological agriculture systems and for extending ecological agriculture technology. In 1994, the Ministry of Agriculture, together with other six ministries and commissions, put forward ‘‘Report on developing ecological agriculture’’ sanctioned by the State Council, and required governments at levels of province, region, and county to actively carry on pilot projects on ecological agriculture in accordance with local conditions; they also listed ecological agriculture extension and ecological environment protection as important agenda items. In ‘‘Ninth five‐year plan of the national economy and social development of the People’s Republic of China, the perspective goal programme of 2010, and China’s agenda of the 21st century,’’ the protection of the environment and developing ecological agriculture vigorously were again stressed (Wang, 1998). With the growth of ecological agriculture, many theoretical books about ecological agriculture have been published. In the 1980s, Q. Ye, a well‐known scholar of agricultural economic management and ecological economics, and one of the founders of CEA, published an important book setting out the basic theories (Ye, 1988). At about the same time, Ma and Li (1987) also published a systematic statement of the theory and methods of ecological agriculture. Subsequently, in the light of sustainable development theory, more literature on ecological agriculture has come out, including Yan (2003), Li (2003), Lu (2002), Wang (2001), and Sun (1993). Similarly, some relevant journals started publication, such as ‘‘Agricultural Modernization Research’’ (1980),

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‘‘Contemporary Eco‐Agriculture’’ (1992), and ‘‘Chinese Journal of Eco‐Agriculture’’ (old name: ‘‘Eco‐Agriculture Research,’’ 1993). The Chinese Ecological Economics Society (CEES) was founded in Beijing in 1984 and is now the biggest institution for systematically studying the harmonious interaction between ecology and the economy.

IV. PRACTICAL ASPECTS OF CEA CEA inherits the quintessence of Chinese traditional agriculture. It takes in modern agricultural science and technology, fully utilizes but protects natural resources, and eVectively connects diverse agricultural components, including farming, forestry, animal husbandry, sideline production, and fisheries. The main practical aspects of CEA are summarized as follows (Cao et al., 2002; Jin and Jin, 1999; Luo, 1994; Sun et al., 1990; Tao, 1993).

A. VERTICALLY DISTRIBUTED FARMING Within a natural biotic community, organisms in diVerent layers each have their own ecological niche and mutualistic symbiosis. In this way, they fully exploit environmental resources (such as air, sunlight, and the like). Applied to agricultural production, these characteristics and relations are used to establish a multilevel structure in space and multisequences in time in order to fully utilize space, raise solar energy use eYciency, and boost land productivity. Typical applications of vertically distributed farming technology include intercropping (forest–grain, forest–herb, fruit–grain), mixed cropping, under‐crop sowing, rice–fish symbiosis, combining forest, shrubs, and grasses, and multilayer stocking of water bodies. A successful case of the application of this technique is an artificial community of tea tree and gum tree in Xishuangbanna, Yunnan province. According to the stratification of forest community, this artificial community was designed with gum trees at the highest layer (5–6 m), cinnamon and rauwolfia at 3–4 m, large‐leaved tea at about 1 m, and finally hygrophilous and shade‐tolerant amomum fruit, a valuable drug. These multiple layers promote ecological harmony (Zheng et al., 1994).

B. MULTILEVEL ORGANIC SUBSTANCE UTILIZATION Organic substances can be used at several levels to mimic the food chain structure and create an artificial ecosystem. The output of one subsystem (excrement) is the input of another, following which the excrement is used again in another production process, forming a stable material cycle. This makes

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full use of natural resources, increases production, reduces pollution, and gives economic benefits. For example, in some ecological farms, chicken excrement is fed to pigs, pig excrement is fed to fish (or enters the methane pool), and the mud in the fishpond (or sludge from methane fermentation) is used as a crop fertilizer. In turn, crop products are foraged by chickens and pigs, thus forming a benign material cycle. Such system‐wide farming provides an eYcient integration of primary agriculture and sideline products (Zheng et al., 1994). A traditional ecological agriculture pattern of mulberry trees and fishponds in fact dates back to the Ming Dynasty. It involves digging a pond and raising the base land, following the natural topography, so that the area of fishpond and land are broadly comparable. People use sludge in the pond to fertilize the mulberry trees, feed mulberry leaves to silkworms, and then feed the silkworm excrement to fish. The fishpond contained diVerent species (grass carp, bighead, crucian carp, carp, and silver carp, each with diVerent feeding habits), and because they occupy diVerent layers, they form a symbiotic cycle of ‘‘mulberry thick, silkworm fat, and fish strong.’’

C. ENERGY EXPLOITATION An important part of ecological agricultural construction is to supply additional energy to rural areas. New types of energy can be tapped without having adverse eVects on the ecosystem: agricultural wastes can be used for methane fermentation, sunlight can be harnessed for solar cookers and solar water heaters, wind energy can be used for small‐scale power generation, and savings can be made with firewood‐eYcient cooking ovens (Jin and Jin, 1999).

D. INTEGRATED CONTROL TECHNIQUES Integrated techniques for controlling plant diseases and insect pests protect biodiversity and improve the environment. They have been adopted in China in a number of forms to provide high quality, high yield, and security of crops. The wide‐ranging nature of the techniques include using organisms (pests or bacteria) against plant diseases and pests, protecting natural enemies, improving pesticide application technology, carrying out rotation of crops, and selecting pesticides with high eYciency, low toxicity, and low residues (Jin and Jin, 1999).

E. INTRODUCING NEW VARIETIES Empty ecological niches always occur in artificial agricultural ecosystems due to the relatively few species involved—which is the reason why artificial ecosystems are so unstable. Filling up ecological niches involves a combination

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of biotechnology and ecological engineering: biotechnology is applied to genetic resources to select new varieties, with an optimal set of genes, appropriate to chosen ecological niches. Under favorable conditions, productivity can be increased manyfold.

F.

COMPREHENSIVE MANAGEMENT OF AGRICULTURAL ENVIRONMENT

THE

Integrated methods, combining biological and engineering measures, have been introduced to remedy widespread problems in the agricultural environment. They include soil erosion control technology (tree and grass planting and building terraces), desertification control (revegetation), and soil amendment to overcome salinization (combining water conservation and agricultural, chemical, and biological amendment methods). Comprehensive management of agricultural environments has already been applied successfully to control soil erosion in south China and the Loess Plateau, salinization in the Huang–Huai–Hai basins of the North China Plain, and desertification in the northwest (Luo, 1994).

V. CASE STUDIES OF CEA A. ECOLOGICAL AGRICULTURE

IN

MOUNTAINOUS REGIONS

The foothills and mountainous regions of China occupy about 70% of the country. Sometimes the rugged terrain severely limits agriculture and often makes communication diYcult. On the other hand, the complex mixture of ecosystem types and the rich diversity of species and natural resources can be seen as definite advantages. If used well, an impoverished situation can be changed into a highly productive one. Therefore, developing distinctive ecological agriculture in tune with local conditions can greatly assist sustainable development. The worst ecological problem in the foothills and mountains is soil erosion, mainly caused by vegetation deterioration. Starting with planting trees on hillsides for water and soil conservation, a diversified complex agricultural economy can be built, one that includes stockbreeding, forestry, fruit trees, and local products industries (including cottage crafts). The construction of ecological agriculture in Wulian county of Shandong province, East China is a good example.

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1. General Situation of Wulian County Wulian county is located at the southeast of Shandong Peninsula. With a warm temperate and semi‐humid climate, Wulian’s area of foothills and mountains accounts for 94% of the whole county, and the cultivated land per capita is only 0.067 ha. It is short of water resources, the distribution of precipitation is uneven, and groundwater is scarce. Before 1978, agricultural production in Wulian grew quickly, but because of a lack of ecological consciousness, land cover was greatly changed at the expense of damaging forest and grassland, resulting in ecological deterioration.

2. Ecological Agriculture Construction in Wulian County The process of ecological agriculture construction in Wulian county can be divided into three stages. (1) In the early stage, the 1970s and 1980s, the kernel of ecological agriculture began with a comprehensive management plan combining biological measures with engineering on a small catchment scale. Starting with separate farm, forestry, and grassland industries, a complex ecosystem was built which intermixed farming, forestry, and animal husbandry to give a more diversified economy. When revegetation was done, great attention was needed to establish two main systems: ecological forest and production forest. In the first, major eVort was devoted to developing the forest for water and soil conservation in remote mountain areas, taking pine and locust tree as key species; shelter belts were also set up in hilly land and on drainage lines, taking willow as the key species. Once established, the way was open for developing economic (production) forests in submountainous regions, combining trees, shrubs, and grasses in order to improve shielding eVects. Through such measures, field climate was improved, soil erosion was controlled, and farmer income increased. An adequate supply of rural energy is very important in the initial stages of ecological agriculture. Ways of doing this include methane digestion, fuel‐ saving stoves, and solar water heaters. Rural energy supplies can improve material cycling, and hence the environment, and can increase profits. (2) In the 1990s, ecological agriculture in Wulian grew quickly. The main achievements at this stage included the following aspects: (A) Comprehensive control on mountain, water, forest, and field increased the eYciency of natural resource use. Previously, where forest had been used only for water and soil conservation, it was replaced by an eco‐forestry system combining ecological benefits with profit‐making forestry. The system typically used a layout of pine and locust tree at the top of a mountain, chestnut on the

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slopes, apple, pear, and peach below, and grains and oilseed plants at the foot. (B) Primary productivity and input eYciency were increased by further adjusting the overall structure. Practical measures were to conduct vertical planting and extend intercropping, undertake plastic mulching, and popularize integrated control methods of pest management. The consumption of organic fertilizer was increased through techniques such as formula fertilization, returning straw to the field, and transferring city sewerage sludge to the countryside to improve soil fertility. (C) To strengthen secondary production, scale‐up of farming, forestry, stockbreeding, fisheries, and processing industries was encouraged. In particular, stockbreeding could lead to products such as meat, eggs, milk, and so on. Hay and straw could also be used to improve biological conversion eYciency and land use capability; they also provide a source of organic fertilizer for crops, trees, and fruits. (3) From the end of 1990s until now, several problems associated with the development of township enterprises arose. They included environmental pollution, loss of labor force, lack of capital, and transfers of land use rights. Nevertheless, through further popularizing biological and integrated control, increasing organic fertilizer inputs, developing premium‐value ‘‘green’’ food and organic food, and carrying out water‐saving agriculture, a pattern has been built up of agricultural production with intensive, highly eYcient, and sustainable features. Furthermore, ecological agriculture has already begun moving toward township enterprises, which develop local processing, transport, and sale of agricultural and sideline products. The benefits of ecological agriculture in Wulian county are plain at the ecological, economic, and social levels. Compared to 1993, in 1997 the amount of organic fertilizer application increased by 44.4%, forest coverage reached 50.4%, gross fruit yield increased by 35%, and the area of soil erosion was reduced by 74%. The average growth rates per year of gross national product, gross national product per capita, and net income per capita were 35.7, 35.3, and 43%, respectively. The gross value of stockbreeding represented 34.5% of total agricultural output in 1997, with a growth rate per year of 53%. Commodity prices for agricultural products rose by 64% over the same period.

B. WATER‐COLLECTING ECOLOGICAL AGRICULTURE IN WESTERN CHINA Regions in the west of China, taking in 71% of the country, are rich in natural resources, although largely underexploited. The area possesses numerous mountain ranges and rivers and a unique, but vulnerable environment. The major environmental problems are vegetation loss, land degradation due to soil erosion, desertification, and salinization. Nearly all of these problems are related to water, so the lack of this resource is

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the bottleneck restricting social and economic development. In western China, the main industries are traditional agriculture and stockbreeding, and the agriculture‐dependent population accounts for 80% of the total. Sustainable agricultural development therefore plays a very important role in the region. Annual precipitation in northwest China is only 200–400 mm, and it is concentrated in July to September as heavy storms. Precipitation often cannot satisfy the water requirements of summer‐harvested crops such as wheat and pea, the main crops in the region, resulting in low and variable yields. Although rainfall is the main source of water for agriculture, rainwater use eYciency was still low at only 14–32% (Shang and Chang, 1999). Developing better ways of collecting rainwater is therefore of great significance in arid areas. There has been a trend in mountainous areas from traditional rain‐fed agriculture to rainwater‐harvesting agriculture (Zhang, 1999). In recent years, a catchment system for better rainwater utilization has been successfully developed for both domestic water use and agricultural production in arid and semiarid regions, especially in the poorest mountain areas in northwest China. Research on water‐collecting ecological agriculture began in the middle of 1980s and was based on agricultural practices in arid and semiarid regions of the Loess Plateau. The central target of water‐collecting ecological agriculture is to set up a harmonious agricultural–ecological–economic system in which the runoV generated from artificial or natural catchment surfaces is collected by storing it in special water storages and is then used for necessarily limited irrigation. It is a sustainable development pattern that simultaneously gives consideration to economic, ecological, and social benefits. The system of water collection is basically made up of three subsystems. (1) An engineering subsystem for water collection. This collects rainwater falling on hard surfaces like courtyards, roofs, and stores it in tanks and ponds. (2) An agronomic engineering subsystem. The function here is to produce various agricultural products using a range of complementary water‐saving irrigation schemes. Drip irrigation, moving drip irrigation, small‐scale sprinkling irrigation, drip irrigation under mulching film, and drip irrigation in large‐scale greenhouses had all been used (Deng et al., 2000; Yin et al., 2000). The combination of rainwater catchment and complementary irrigation at critical stages of crop growth can increase both crop yield and water use eYciency. This method is an eVective way to overcome water shortages in semiarid and arid regions and to promote sustainable development of dryland agriculture. (3) A subsystem for social economy and management. In the northwest, people’s consciousness of how to market a commodity is low. Organizing workers to market agricultural products is a major step in moving from a small‐scale worker economy to a commodity‐based economy.

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Water‐collecting ecological agriculture can significantly improve agricultural productivity, optimize agricultural structure, rehabilitate and rebuild the ecological environment, and improve a region’s resistance to natural risks, as the following example demonstrates. In Lijiajian village, Yuzhong county, Gansu province, 51 water cisterns have already been built, 17 of them for drinking, another 34 for supplementary irrigation. With water‐ collecting ecological agriculture, new crop‐planting patterns, new techniques of water‐saving irrigation, and large‐scale greenhouse cultivation, the yield reached 166.7–233.3 kg hm 2 and the average per capita income more than US$360, a record for arid areas (Yu and Chen, 2001).

C. FROM ECOLOGICAL AGRICULTURE

TO

ECOLOGICAL INDUSTRY

Over the last 20 years, CEA has achieved much, both in technology and in management of vast rural areas. Nevertheless, it should be pointed out that CEA has remained at low levels of technology, profit, and scale, and there is still a big gap to industrialization scale and agricultural modernization. Indeed, in the rich coastal areas ecological agriculture is shrinking. Sustainable development will only succeed in the rural areas of China if the small agricultural cycle is replaced by a larger industrial cycle combining industry, agriculture, and business. In this way, the small‐scale farmer economy will be replaced by a commodity economy so that ecological agriculture in China will be transformed into ecological industry (Wang and Jiang, 1991). Ecological industry is organized according to the principles and laws of ecological economics, and is based on ecosystem carrying capacity. A production system is set up to imitate the structure and processes of the natural ecosystem. Many industries are combined to provide the full recycling of material without any waste, there is multilevel utilization of energy, and environmental impacts are minimized. Instructed by ideas in ecological engineering, ecological industry not only emphasizes the rational connection between components, but also stresses no‐waste resource recycling. In short, it seeks the harmonious development of ecology, economics, and society. On the basis of the rich experience of ecological agriculture over the last 20 years, it rationally couples farming, livestock breeding, processing, industry, and service trades. Lu and Zhao (2001) believe that constructing an ecological garden industry might become one of the major achievements of sustainable development in China. A good example of ecological industry is the system of sugar‐refining/ alcohol/energy/agriculture in Hainan province. Here, sugar‐refining is the pillar of industry in the region, which is well suited to growing sugarcane. However, liquid wastes from sugar‐refining have become a severe problem, preventing the development of the industry. The waste has a high organic matter concentration, and is diYcult and costly to treat.

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The approach of the ecological industry pattern is to combine pollution control with integrated utilization of resources. This maximizes the economic benefits and improves eYciency of the production system, while minimizing environmental impacts. Starting with the integrated utilization of waste liquid from sugar‐refining, the production is designed stepwise, organically linking the sugarcane growing with sugar‐refining, alcohol production from waste liquid, methane production from alcohol waste, and electricity generation from methane. In turn, liquid waste from methane production is used on farmland as organic fertilizer, and solid waste from the electric power plant (fly ash and slurry) is used to produce building materials. The industrial chain thereby makes use of the productive potential of all the resources. Ouyang et al. (2004) set out details of how the system works. Another example is Liuminying ecological farm, 25‐km southeast of Beijing city. It is located at Liuminying village, Zhangziying town, Daxing county, on the alluvial plain of the Yongding river. It is in the warm temperate and semi‐ humid climate zone and is underlain with abundant underground water with the table at a depth of about 1.5 m. The soil organic matter content is about 1.6%, giving a high ability for conserving soil fertility (Zhang et al., 1993). Before 1982, the agricultural production structure was very simple. Some 95% of the labor force was involved in crop production, the output of which accounted for 78% of total earnings. Crop yield relied mainly on the input of chemical fertilizer, which resulted in soil sealing and environmental pollution. Therefore, the first change was to broaden activities to take in farming, forestry, animal husbandry, sideline production, and fish culture. To maintain grain production, 26.7 hm2 were set aside for new vegetable greenhouses and 20 hm2 for an orchard and a seedling nursery. For livestock breeding, five modern animal farms and a fish pond of 4 hm2 were established, and these provided plentiful supplies of organic fertilizer for crop cultivation, sharply reducing the need for chemical fertilizers. At the same time, to diversify the village economy, many pollution free and high economic benefit enterprises were set up, such as a feed processing plant, food and drink factory, small‐scale butchery, machine‐ making factory, and handcraft works. These provided products to Beijing, and brought money to the farmers (Zhang et al., 1993). The availability of energy is an important factor aVecting agricultural production. Ecological agriculture aims to be as energy self‐suYcient as possible, mainly through use of solar and biological energy, especially by the integrated utilization of organic waste from the agricultural ecosystem. Working by thermophilic fermentation, two 100‐m3 methane‐generating stations were established (Fig. 3). Wastes discharged from farm houses and farms—such as excrement from chickens, cows, and pigs—contain protein, fat, and other nutrients, so after purification the sludge generated from methane fermentation can be used as an organic feedstuV for fish, and an organic fertilizer on farmland, vegetable

ECOLOGICAL AGRICULTURE IN CHINA Chicken, duck, yoghourt, meat

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Energy for dining room and factory Feed cycle Fertilizer

Chicken farm

sh ar M as g

Market of meat egg milk fish

Cattle farm

Pig farm

Fish pond

Corn stalk

Fertilizer Food factory processing Pou Marsh gas idue dun ltry s Re id g Cow liqsuidue e R dung liquid Marsh gas Resid station ue Pig liquid manure Re s id ue sid Marsh gas manure liquid ue Re id il qu Farmer

Farm land

Vegetable land

Orchard

Market of grain vegetables fruits seedling tree

Seedling nursery

Feed cycle

Figure 3 The ecological agriculture pattern linked by methane (Liuminying ecological farm). Adapted from http://www.liuminying.com.cn

gardens, orchards, and seedling nurseries; it can also be used as a feed additive or in making a biochemical pesticide. The liquid residue can be used as animal feed and field manure, or it can be used for drip irrigation in greenhouses. Sprayed on a leaf surface, the liquid acts as a fertilizer, reducing the need for chemical fertilizers and pesticides; the liquid also increases soil organic matter and improves yields of crops, meat, egg, and fish. Gas generated by the methane station can be used directly for household use or energy for the small local industries. The gas replaces burning of crop straw and provides a valuable use of animal excrement, in the process reducing the likelihood of pollution. After 1985, the landscape planning engineering was carried out at Liuminying ecological farm. Many trees, including fruit trees, were planted, and the percentage of forest cover increased from 6 to 20%, improving the field microclimate. At the same time, in cooperation with the Chinese Academy of Forestry Sciences, vertical planting and aquaculture were brought in, imitating natural ecosystem. The aim was to imitate natural ecosystems and establish a mutualism relationship among organisms. This improves space resources, and promotes material recycling and full energy utilization; it also uses integrated pest control to prevent harmful organisms and materials entering the ecosystem. In an example of intercropping, a 20‐hm2 paulownia–cereal plot and a 13.3‐hm2 paulownia–jujube–cereal plot were introduced, each increasing crop yield, income, and green food diversity. A new point of economic growth of the farm is to introduce sightseeing agriculture and ecological tourism. In 2005, the town government of Zhangziying invested about 2.5 million US$ to establish a 10‐km ‘‘Vegetable Avenue.’’ In May of this year, a water‐saving pipeline was laid. If all goes

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well, vegetable supply should increase by some 27 million kg year 1, which includes 4 million kg of out of season vegetables. ‘‘Vegetable Avenue’’ will be the largest single scenic spot in Beijing, a mixture of vegetable production and selling, technical development, new variety trials, and, of course, sightseeing and picking.

VI. PROBLEMS It has been demonstrated that ecological agriculture can avoid the disadvantages brought by petrochemical agriculture. Moreover, it can fully and eVectively make use of natural resources, thereby raising productivity, maintaining the natural ecological balance, protecting the environment, and improving the recycling of material and energy. The result is a graceful, relaxed, civilized, and highly functional living environment. Developing ecological agriculture is essential for achieving agricultural sustainable development in China. In the past 25 years, CEA has developed rapidly in size and diversity, and it has provided enormous social, economic, and ecological benefits. Nevertheless, over this time, the demonstration area has only reached 7% of the cultivated land (Jin and Jin, 1999). Many diYculties and impediments still exist for full implementation of ecological agriculture over vast areas of rural China (Li and Luo, 2001; Shen et al., 2002; Wang, 1993; Wu et al., 2001; Yang, 2004).

A.

FROM THEORY TO PRACTICE

Ecological agriculture is part of a bigger picture: the whole mega‐agriculture of farming, forestry, animal husbandry, sideline production, and fish culture. The old idea of high‐price products, low‐price resources, and the environment of no value is still deep‐rooted among many. An important factor is to motivate the whole community, right through from workers to local governments, always aiming to resolve the inherent conflict between overall and long‐term ecological benefits and partial and short‐term economic benefits. Policies, laws, and regulations for promoting and protecting ecological agriculture still need to be perfected in China.

B. POVERTY A number of rural areas face multiple and persistent problems: overpopulation and surplus labor, shortage of natural resources, serious land degradation, grassland degradation, and desertification. The agricultural infrastructure is

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often poor, so many regions fail to escape the vicious cycle of population increase, environmental damage, and slow economic growth.

C. LACK

OF

SYSTEMATIC THEORETICAL RESEARCH

During more than 20 years’ development, more attention has been paid to the application and practice of ecological agriculture, while deep exploration has not been carried through at the theoretical level. It lacks a systematic base and improvements in economics, ecology, and systems theory are needed.

D. SMALL PRODUCTION SCALE CEA deals mostly with a small unit: a farming family or a village. However, the county is the basic unit of Chinese administrative system, with relatively independent administration and economic autonomy. The construction of CEA should therefore increase in scale from the level of a farming family or a village to the county level.

E. LACK

OF

FUNDS

In China, the shortage of funds is a great diYculty for CEA implementation. The economic base is so weak that it holds back the basic engineering construction of ecological agriculture. Financial investment in agriculture is seriously insuYcient: investment for agricultural research only accounts for 0.5% of annual gross national product, while in developed countries this ratio approaches 5% or more.

F. EDUCATION

OF

FARMERS

Whether the construction of ecological agriculture succeeds or not, the key problem is the degree to which farmers accept the ideas of ecological agriculture. In production practice, short‐term economic gains and long‐term environmental benefits need to be balanced. Chinese farmers have low levels of education, poor technical skills, and very little knowledge and experience of ecological agriculture. These factors hinder the large‐scale implementation of ecological agriculture. In addition, due to low returns from agricultural activity, talented young adults move away from the countryside, slowing the uptake of ecological agriculture.

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G. LACK

OF

MARKET COMPETIVENESS

Because the level of agricultural modernization is relatively low in China at present, the future for CEA lies in actively pushing toward industrialization. However, most agricultural products presently lack market competitiveness because of old production techniques and substandard quality of agricultural products. Furthermore, most farmers lack a clear understanding of the market economy. Laws, policies, systems, and management measures conducive to ecological agriculture need to be established and perfected before we begin to see raised production standards, pure residue‐free products, and eYcient management.

VII. CONCLUDING REMARKS Entering the twenty‐first century and facing increasing globalization and great pressures for environmental protection, CEA encounters new opportunities and challenges. We believe it should try to achieve three things: a change from pursuing the amount of products to stressing its quality, from seeing one domestic market to considering two markets of home and abroad, and diversifying from a single productive function to multiple productive and ecological functions. In order to successfully realize these changes, great attention should be paid to the following aspects (Guo, 2004; Jin and Jin, 1999; Li, 2004; Sun et al., 1993).

A. FURTHER STUDIES

ON

THEORIES

OF

CEA

The rich knowledge of ecological agriculture acquired over the last 20 years should be thoroughly and systematically summarized, and existing knowledge should be placed in a theoretical context. The concepts and objectives of CEA, the definitions of ecosystems, and the basic theory of ecological agriculture should be improved. The basic theory relates to the theories of energy flow and material cycles, of ecological niches, and of biodiversity, stability, and sustainability. Optimization of ecological agriculture should be achieved according to ecological engineering principles, and then its powerful productivity gains could be actualized.

B.

MODERN TECHNIQUES APPLICATION

The goals of ecological agriculture are high yield, top quality, high eYciency, low consumption, and no pollution. These could be realized through increasing use of modern techniques derived from traditional agriculture technology. The

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techniques could include methods of agricultural biotechnology such as biogenetic engineering, enzyme engineering, and microorganism utilization. Other possibilities are techniques of ecological engineering as used in environmental protection: recycled energy engineering, waste reclamation engineering, integrated pest control technology, and bioactive fertilizer utilization. Finally, benefits may flow from using geographic information systems, remote sensing, and global positioning systems at the macroscale level.

C. ORGANIC FOOD: FILLING

A

GAP

IN THE

MARKET

With the rise in people’s living standards, demand for pollution‐free ecological food is growing, particularly on the international market. Safe food is the product of highly eYcient ecological agriculture, and is marketed as ‘‘green’’ or organic food. Developing pollution‐free food is an arduous task and needs careful attention to the input of chemical products, especially nonacceptable input materials. Integration of agriculture, industry, and commerce could take ecological agriculture in the direction of agricultural industrialization.

D. ECOLOGICAL AGRICULTURE

AND

TOWNSHIP ENTERPRISE

In China, the township enterprise has played an important role in the transition of rural economy from traditional agriculture to the industrialized kind, but at the same time it has raised a series of rural environmental issues that cannot be ignored. Ecological agriculture emphasizes the rational utilization of natural resources, recycling of materials, and waste reclamation, and in this way reduces pollution brought about by township industries or agriculture itself. The township enterprise can reasonably provide funds for the application of ecological agriculture technology, and so the two enterprises can depend on and promote each other.

E. INTERNATIONAL COOPERATION

IN

ECOLOGICAL AGRICULTURE

Exchanges with foreign countries on their experiences of ecological agriculture are presently few and far between. International cooperation on ecological agriculture would lead to information exchange, staV training, symposia, cooperative research and development, benefiting all sides. In conclusion, CEA has made impressive progress over the last 20 years, but there are major problems to be overcome before it can fulfill its promise. Ingenuity and the dedication of many Chinese agricultural scientists will be needed in order to make major progress.

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ACKNOWLEDGMENTS This chapter was supported by NSFC (National Natural Science Foundation of China) project 40275004 and the opening fund project of Key Laboratory for Water‐saving Agriculture of Hebei Province (0508021‐HBKLA‐03).

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Liu, X. (1995). On the sustainable development of Chinese agriculture in the 21st century. J. Nat. Resour. 10(3), 216–224. Liu, C., and You, M. (1993). Sustainable development of agriculture in the North China Plain and south‐to‐north water transfer. Eco‐Agric. Res. 1(1), 45–50. Liu, Y., and Wu, C. (2000). Advances and theory of agricultural sustainable development. Econ. Geogr. 20(1), 63–67. Liu, Y., and Wu, C. (2001). Typical models and approach of sustainable agricultural development in foreign and home. J. Nanjing Normal Univ. 24(2), 119–124. Lockeretz, W. (1988). Open questions in sustainable agriculture. Am. J. Altern. Agric. 3(4), 174–181. Lockeretz, W., Shearer, G., and Kohl, H. D. (1981). Organic farming in the corn belt. Science 211, 540–547. Lu, L. (1995). China’s sustainable agricultural development. China Popul., Resour. Dev. 5(2), 27–33. Lu, M. (Ed.) (2002). ‘‘Modern Ecological Agriculture.’’ China Agriculture Press, Beijing. Luo, S. (1994). The technical system of ecological agriculture in China. In ‘‘Proceedings of the Symposium of Environment and Development of China in the 21st Century’’ (W. Niu and H. Yu, Eds.), pp. 391–397. Lu, B., and Zhao, J. (2001). Ecological industry garden: An ideal model of sustainable development. Environ. Sci. 22(2), 1–6. Ma, S. (1983). Ecological engineering—the application of ecosystem principle. Acta Ecologica Sinica 4, 20–22. Ma, S., and Li, S. (Eds.) (1987). ‘‘Ecological Agriculture Engineering in China.’’ Science Press, Beijing. Merrill, M. C. (1983). Eco‐agriculture: A review of its history and philosophy. Biol. Agric. Hortic. 1(3), 181–210. Ouyang, Z., Zhao, T., Miao, H., Wang, R., and Wang, X. (2004). Design for ecological industrial chain for sugar refining, alcohol distillation, energy provision and agriculture in Hainan. Acta Scientiae Circumstantiae 24(5), 915–921. Scofield, A. M. (1986). Organic farming: The origin of the name. Biol. Agric. Hortic. 4(1), 1–5. Shang, X., and Chang, J. (1999). Storage and utilization of rainwater and development of rural economy in central Gansu. Agric. Res. Arid Areas 17(2), 116–121. Shen, H., Kang, X., and Zhang, W. (1993). Chinese agricultural modernization and the analysis of the eco‐economy at its different developing stage. Eco‐Agric. Res. 1(2), 15–26. Shen, M., Liang, Y., and Gao, G. (2002). A cogitation upon China’s sustainable agriculture development. Green Econ. 5, 59–61. Sun, H. (Ed.) (1993). ‘‘Theory and Method of Ecological Agriculture.’’ Shandong Science and Technology Press, Jinan. Sun, H., and Hu, T. (1993). The prospect of eco‐agricultural development in China in the current movement of the world sustainable agriculture. Eco‐Agric. Res. 1(1), 25–30. Sun, H., Han, C., and Zhang, R. (1990). Characteristics, principles and major techniques of Chinese ecological agriculture. Res. Agric. Mod. 11(3), 3–8. Sun, H., Hu, T., and Zhang, R. (1993). The prospect of eco‐agricultural development in China in the current movement of the world sustainable agriculture. Eco‐Agric. Res. 1(1), 25–30. Tao, S. (1993). The ecological agriculture and its development in China. Eco‐Agric. Res. 1(2), 1–5. Walter, G., and Young, D. L. (1987). An agronomic and economic comparison of a low‐input cropping system in the Palouse. Am. J. Altern. Agric. 2(2), 51–56. Wang, H. (1993). A study on the obstructions, regularity and countermeasures of sustainable agricultural development in China. Eco‐Agric. Res. 1(2), 31–39. Wang, W. (1998). Theory and practice of eco‐agriculture in China. Environ. Her. 2, 5–8.

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COTTON LEAF PHOTOSYNTHESIS CARBON METABOLISM

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W. T. Pettigrew1 and T. J. Gerik2 1 United States Department of Agriculture, Agricultural Research Service, Crop Genetics and Production Research Unit, Stoneville, Mississippi 38776 2 Blackland Research Center, Temple, Texas 76502

I. Introduction II. Genetic Variability A. CO2 Exchange Rate and Stomatal Conductance B. Chlorophyll Fluorescence C. Photosynthetic Enzymes D. A:Ci Curves E. 13C Discrimination III. Management and Environmental EVects A. Plant Growth Regulators B. Plant Nutrition and Soil Fertility C. Moisture Stress D. Temperature IV. Conclusions V. Summary References

Photosynthesis is the basis of plant dry matter production and a major determination of yield in cotton (Gossypium hirsutum L.). Much of the cotton yield increases in recent years can be attributed to the improved partitioning of dry matter into reproductive growth rather than vegetative growth. However, this strategy can only be taken so far before the amount of photosynthesizing leaf area becomes the limiting factor. Therefore, improved plant photosynthesis coupled with good dry matter partitioning could lead to additional yield improvements. Research has identified both genetic and environmental variations in the rate of cotton photosynthesis. Superior leaf photosynthetic performance has been exhibited by okra and super‐okra leaf types compared to the normal leaf types. Photosynthetic variation has also been identified within the normal leaf type pool of germplasm. However, geneticists have generally not targeted this trait for genetic improvement in cotton. In addition, leaf tissue concentration of the three major plant nutrients (nitrogen, potassium, and phosphorus) need to be maintained at suYcient levels for optimum photosynthesis. Under deficient soil fertility conditions, supplemental fertilization can increase overall growth due to 209 Advances in Agronomy, Volume 94 0065-2113/07 $35.00 DOI: 10.1016/S0065-2113(06)94005-X

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W. T. PETTIGREW AND T. J. GERIK both increased leaf area production and increased photosynthetic rate per unit leaf area. Both excessive and deficient soil moisture conditions can depress the photosynthetic performance of the plant and its corresponding growth. Similarly, an optimum temperature range exists, above and below which the photosynthesis is negatively impacted. This knowledge of variation in both genetic and environmental influences on photosynthesis oVers hope of improved photosynthetic performance through either a concerted genetic selection or modified production systems that minimize exposure to some of the rate‐limiting environmental conditions.

I. INTRODUCTION After an impressive history of almost continual yield increases, the yield potential of modern cotton genotypes appeared to have plateaued in the mid‐ 1990s relative to the historical yield performance of older genotypes (Meredith, 1995). Although improved cultural practices can account for much of the increase, the genetic component of this yield increase has primarily come from increasing the amount of dry matter partitioned into reproductive growth instead of vegetative growth, that is, harvest index (Wells and Meredith, 1984). There is a limit as to how far this breeding strategy can be utilized before leaf area becomes limiting. To forestall this limitation, we must strive to make the most eYcient use of this stable or declining leaf area by optimizing photosynthetic production. The prospect of improving photosynthesis can be usefully approached from two diVerent aspects: (1) genetic variability and (2) management practices that aVect plant growth or mitigate negative environmental eVects.

II. GENETIC VARIABILITY A. CO2 EXCHANGE RATE

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STOMATAL CONDUCTANCE

Photosynthesis is the basis of dry matter production and the major factor in yield determination. The overall, multistep process is highly complex, with many points of regulation in its pathway, the end result of which can be measured as the rate of decline in ambient CO2 concentration when a leaf is enclosed in a clear, sunlit chamber. The multicomponent nature of photosynthesis implies numerous points where one could intervene genetically and thereby possibly influence the overall rate or eYciency of photosynthate production. Some initial work addressing the genetic component of photosynthesis in cotton involved leaf gas exchange diVerences among Gossypium species. For all the gas exchange studies

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reported in this chapter, unless other conditions are specifically mentioned, the measurements were conducted on single leaves under light levels presumed to be saturating [1600 mmol m2 s2 photosynthetic photon flux density (PPFD) or greater]. El‐Sharkawy et al. (1965) performed a survey on 26 species of Gossypium and found the leaf CO2‐exchange rates (CER) to vary from a high of 25 mmol m2 s1 to a low of 11 mmol m2 s1 among species. Muramoto et al. (1965) detected no measurable variation in leaf CER among genotypes of G. hirsutum and G. barbadense or among F1 crosses between the two species. On the other hand, Elmore et al. (1967) found significant leaf CER diVerences among several races of G. hirsutum when compared with cultivated varieties of G. hirsutum and G. barbadense. These authors also found that an interspecific hybrid between G. hirsutum and G. barbadense produced greater leaf CER than that of its parents. Other than the possible improved photosynthesis of an F1 interspecific cross, this early work did not suggest that genetic improvement in G. hirsutum or G. barbadense leaf photosynthesis could be observed and exploited in other Gossypium species. A number of unusual pigment and leaf type variants exists among G. hirsutum, providing useful germplasm for photosynthetic research in cotton. Gas exchange properties of a chlorophyll‐deficient virescent mutant of upland cotton (G. hirsutum) were compared with those of a normal pigmented wild‐type upland cotton genotype by Benedict et al. (1972). Under high light conditions (230 W m2), they reported similar rates of CER between the wild‐ type and virescent cotton lines on a leaf area basis, but the virescent mutant demonstrated increased CER compared to the wild‐type on a chlorophyll basis. In contrast, Muramoto et al. (1965) reported that a Sea Island Virescent variety of a Pima cotton (G. barbadense) had lower chlorophyll concentration and lower CER on a leaf area basis than a comparable normal pigmented Pima variety. Considerable research has been performed on leaf shape variants of G. hirsutum. These leaf shapes can vary from the highly clefted and relatively small leaf area of the super‐okra leaf type to the normal clefting and leaf area of typical leaf types used in US cotton production. Okra and sub‐okra leaf types are intermediate between super‐okra and normal leaf types in their degree of clefting and leaf area reduction. The degree of leaf clefting and area reduction associated with these leaf shapes can also vary depending on the genetic background in which the leaf shape trait is expressed. Gas exchange research on leaf shape variants over the years has produced inconsistent results. Elmore et al. (1967) found similar leaf CER between super‐okra and normal leaf. Okra, super‐okra, laciniate, and normal leaf type isolines did not diVer significantly in 14CO2 fixation for individual leaves within the canopy (Karami et al., 1980; Kerby et al., 1980) or canopy CER when expressed on a leaf area basis (Pegelow et al., 1977), although in each case there was a trend for the laciniate, super‐okra, or okra isolines to have greater leaf photosynthesis. When expressed on a ground area basis,

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Pegelow et al. (1977) and Wells et al. (1986) found lower canopy CER for okra leaf which they attributed to insuYcient leaf area. Sub‐okra leaf type, however, demonstrated similar canopy CER on a ground area basis to that of normal leaf. Wells et al. (1986) also reported similar leaf CER for okra and normal isolines of the genotype ‘‘MD 65–11,’’ although the CER of super‐okra was lower than either okra or normal leaf types. In contrast, Peng and Krieg (1991) reported canopy photosynthesis per unit leaf area of okra leaf plants to be greater than for normal leaf plants. Perry et al. (1983) also reported instances where both okra and super‐okra leaf types produced greater leaf CER than their normal leaf counterpart. Using the same isolines employed by Wells et al. (1986), Pettigrew et al. (1993b) found that leaf CER of super‐okra and okra was 24 and 22% greater than the normal leaf isoline, respectively. Furthermore, the elevated CER of okra leaf type lines over that of normal leaf type lines was maintained across multiple light levels used in generating light response curves (Pettigrew, 2004b). Stiller et al. (2005) did not find okra leaf varieties to have greater net photosynthesis rates than normal varieties, but they did not make comparisons among genetic leaf type isolines. It is not clear why Wells et al. (1986) and Pettigrew et al. (1993b) produced diVerent CER results using the same isolines. Possibly the measurements were taken at diVerent times of day (Section III.D on afternoon decline in photosynthesis in this chapter). Because both Wells et al. (1986) and Pettigrew et al. (1993b) found okra leaf to have a greater specific leaf weight (SLW) than normal leaf, Pettigrew et al. (1993b) concluded that the okra leaf variants had a greater concentration of the photosynthetic apparatus per unit leaf area than the normal leaf isoline. This conclusion was reinforced by finding that the okra isoline of MD 65–11 had thicker leaves and greater concentrations of chlorophyll and soluble protein than the normal leaf. Both Wells et al. (1986) and Pettigrew et al. (1993b) reported lower stomatal conductances (gs) for okra and super‐okra compared to the normal leaf type isoline. Under isothermal conditions, lower gs coupled with greater CER gives the leaves of super‐okra and okra higher water use eYciencies (WUE). In later studies, however, Pettigrew (2004a) was unable to detect gs diVerences between okra and normal leaf type isolines, although a diVerent leaf type isoline pair was used is this later study than in the earlier study. Recent work has investigated the genetic basis for variation in CER that exists in both the G. barbadense (Pima) and G. hirsutum (upland) normal leaf type germplasms. The gas exchange properties of four Pima cotton varieties, one advanced breeding line, and one uncultivated primitive Pima line were investigated by Cornish et al. (1991). Because modern lines had higher CER and gs than older, more primitive Pima lines, they concluded that genetic gains in yield of Pima cotton were accompanied by increases in CER and gs. In follow‐up studies, Radin et al. (1994), Lu et al. (1994), and Lu and Zeiger (1994) suggested that yield improvements in modern Pima lines were

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associated with improved heat tolerance due to superior gs and smaller leaf size. Selecting for greater gs from a segregating population of Pima cotton led to lower leaf temperatures, enhanced fruiting prolificacy during the day’s hottest periods, and improved yield. Because CER showed little sensitivity to temperature in the 23–36 C range while gs increased linearly with temperature in the same range, Lu and Zeiger (1994) suggested that selection pressures for higher CER were not directly advantageous for heat tolerance and improved performance at higher temperatures. Srivastava et al. (1995) showed that the higher gs of the higher yielding Pima lines was associated with greater rates of respiration and proton extrusion in their stomatal guard cells as compared to the low‐yielding, low gs lines. Inheritance of stomatal conductance in Pima cotton appears to be multigenic and may have maternal as well as nuclear components (Percy et al., 1996). Genetic variation in photosynthesis has also been detected in the hirsutum normal leaf type germplasm. Averaged across years, ‘‘Acala SJ‐2’’ had a higher radiation use eYciency (RUE) (1.60 g MJ1 0.06) than ‘‘DPL 50’’ (1.46 g MJ1 0.06) or ‘‘Tamcot CD3H’’ (1.31 g MJ1 0.09) during the reproductive period of growth, but not during vegetative growth (Rosenthal and Gerik, 1991). Quisenberry et al. (1994) found genotypic diVerences in leaf CER among five normal leaf upland cotton genotypes diVering in sink‐ to‐source ratios. Also, as the sink‐to‐source ratio of these genotypes increased, the leaf CER also increased, leading to a highly significant and positive correlation between the two traits. Nonflowering cotton lines, which would also have a low sink‐to‐source ratio, demonstrated lower CER but higher photorespiration than fruiting lines (Perry et al., 1983). Genetic variation in leaf CER and gs was also reported in a pool of 18 normal leaf type‐upland cotton genotypes, representing a range of maturities and adaptation regions by Pettigrew and Meredith (1994). When the single leaf, light saturated, CER values for each genotype taken during the boll filling periods were averaged over 2 years, a significant positive association (r2 ¼ 0.21) was obtained between leaf CER and lint yield. Oftentimes, it has proven diYcult to associate leaf CER and yield (Elmore, 1980), which has been attributed to using a single instantaneous CER measurements to estimate a seasonal phenomenon and a single leaf to estimate a canopy eVect (Zelitch, 1982). By averaging together single leaf measurements taken during boll filling, Pettigrew and Meredith (1994) were able to mimick the seasonal and canopy eVects and thereby establish a relationship between yield and leaf CER. In addition to establishing a connection between lint yield and leaf CER, Pettigrew and Meredith (1994) found that some of the parameters of fiber quality were associated with CER. Fiber micronaire (r ¼ 0.77) and fiber maturity (r ¼ 0.80) were also correlated with leaf CER in 1 year of the study, but no significant associations were detected between either fiber strength or fiber perimeter with CER.

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Genotypic diVerences among the 18 lines used by Pettigrew and Meredith (1994) were also detected for SLW, leaf chlorophyll concentration, and leaf soluble protein concentration. Chlorophyll concentration and SLW of these genotypes were positively correlated with leaf CER, but soluble protein concentration was not. Of the physiological traits measured among these genotypes, regression analysis indicated that gs diVerences contributed the most to changes in CER. Unfortunately, this regression analysis did not prove cause or eVect; was CER responding to gs or was gs responding to CER? Because these genotypes varied in CER and for some of the physiological components thought to be involved in CER, Pettigrew and Meredith (1994) concluded that diVerent factors determining CER may have been involved in the selection process. They further speculated that this explanation raises the possibility of producing additional CER improvement by crossing genotypes with supposedly diVerent mechanisms contributing to high CER.

B. CHLOROPHYLL FLUORESCENCE Chlorophyll fluorescence trace measurements have been associated with activities of various components of the chlorophyll light‐harvesting antenna system, photosynthetic electron transport system, and ATP regeneration complex. Collectively these systems are often referred to as the light reactions of photosynthesis. The ratio of variable chlorophyll fluorescence (Fv) to maximal fluorescence (Fm), (Fv/Fm) is proportional to the quantum yield of photochemistry and is therefore an estimate of the maximum photosystem II (PS II) eYciency. This ratio is one of the most easily obtainable, most widely reported, and most understandable components of chlorophyll fluorescence analysis of crop photosynthetic performance. Although chlorophyll fluorescence theory is well established, the technological advances needed for development of small portable instrumentation that permits the measurement of chlorophyll fluorescence in field situations has only been possible since the early 1990s. Further advances now allow for light‐adapted fluorescence measurements to be taken simultaneously with CER measurements, providing both a measurement of PS II quantum eYciency under light‐adapted conditions and a measurement of photosynthetic electron transport rate. Bjo¨rkman and Demmig (1987) compared chlorophyll fluorescence patterns measured at 77 K for several plant species of diverse origins (two of which were Pima and upland cotton) grown under nonstressed conditions and showed that all these species were relatively consistent in producing an Fv/Fm ratio of 0.83. Fluorescence induction curves were used by Wullschleger et al. (1991) to document photochemical and nonphotochemical quenching in upland cotton leaf, bract, and capsule wall

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tissue. Utilizing various conditions to manipulate leaf starch content in upland cotton plants, Warner and Burke (1993) reported that maximum PS II activity as measured by the Fv/Fm ratio decreased as the leaf starch content increased. An initial screening of six, field‐grown, upland cotton genotypes revealed no significant genotypic variation in Fv/Fm (Pettigrew and Meredith, 1994). In subsequent research, however, the genotype ‘‘Dixie King’’ was found to have a significantly lower Fv/Fm than other upland cotton genotypes during 1 year of a 2‐year study (Pettigrew and Turley, 1998). In addition, although the dark adapted Fv/Fm was no diVerent between okra and normal leaf type genotypes, the okra leaf type lines had 14% greater light‐adapted PS II quantum eYciency and photosynthetic electron transport compared to normal leaf type lines (Pettigrew, 2004a). Furthermore, nonphotochemical quenching was also 11% lower in the okra leaf type lines (Pettigrew, 2004a). Chlorophyll fluorescence, as a tool in the field, is a relatively new method which, to date, has yet to be utilized to the full extent of its capabilities in cotton. As more research is conducted on cotton using these fluorescence techniques, more insights will be revealed about the light harvesting and energy conversion process in cotton and about any potential genetic variation for the reactions involved.

C. PHOTOSYNTHETIC ENZYMES Many enzymes of photosynthesis are associated with the carbon reduction cycle portion of the photosynthetic process, although many are also associated with other components of the photosynthetic process. Cotton contains many secondary phenolic compounds concentrated in glandular structures throughout various tissues. While these compounds are essential in conveying a certain degree of insect resistance to the plant, they have also hindered the study of many of the enzymatic systems in the plant. Tissue grinding required to study the enzymes releases these phenolics and terpenoids from their glandular structures, and these compounds in turn bind with enzymes, rendering them inactive. While this diYculty limited early cotton enzyme work to glandless cotton, research over the years has identified chemical protectants that can be added to the grinding buVers to ensure adequate enzymatic activity during the assays, allowing enzymatic studies of glanded cotton varieties normally used in production. For instance, including borate in the extraction buVer significantly enhanced the recovery of active enzymes (King, 1971) and high‐quality RNA (Wan and Wilkins, 1994) from cotton leaves. Early work by Benedict (1972) and Whelan et al. (1970) demonstrated the presence of several Calvin cycle enzymes in cotton leaves, as well as 25 times more ribulose‐1,5‐bisphosphate carboxylase‐oxygenase (Rubisco) than

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phosphoenolpyruvate carboxylase, consistent with the C3 photosynthesis metabolism of cotton. Benedict et al. (1981) further showed that Rubisco activity in G. hirsutum (AD1) (99 mg CO2 dm2 h1) was more than twofold greater than that of G. davidsonii (D3‐d) (44 mg CO2 dm2 h1). No diVerences in Rubisco content per unit leaf fresh weight were detected among cotton leaf type isolines by Pettigrew et al. (1993b). However, because super‐okra and okra exhibited greater SLW than their normal leaf type counterpart, they speculated that super‐okra and okra might have had more Rubisco on a leaf area basis than the normal leaf type. This scenario parallels the CER diVerences observed between leaf types. An initial, limited screening of six G. hirsutum normal leaf type genotypes revealed significant genotypic diVerences ( p ¼ 0.06) in Rubisco activity on a leaf area basis for 1 year of a 2‐year study (Pettigrew and Turley, 1998). The trend for these genotypes was for the low CER genotypes (i.e., Dixie King) to have lower Rubisco activity than the high CER genotypes (i.e., DES 119). Two enzymes associated with photorespiration, catalase, and hydroxypyruvate reductase, were also assayed for these genotypes. A high CER genotype, ‘‘MD 51 ne,’’ was found to exhibit significantly greater catalase activity on a chlorophyll basis and numerically, though not significantly diVerent, hydroxypyruvate reductase activity on a chlorophyll basis. Perry et al. (1983) previously had documented genotypic variation in photorespiration in cotton, leading them to speculate that photorespiration potentially could be reduced genetically. Much work remains before all the enzymes and proteins involved in photosynthesis for cotton have been properly characterized. Work is needed to determine whether genetic variations exist in either concentration or activity of any of these enzymes.

D.

A:Ci CURVES

Plotting the CER or assimilation rate (A) versus the internal CO2 concentration (Ci) over a range in Ci levels provides more information about the photosynthetic process than ordinary gas exchange measurements, although A:Ci curves can be very labor intensive to generate. Information that can be derived from these curves include CO2 compensation points (an estimate of photorespiration), the maximum assimilation rates at high Ci (Amax) (an indication of eYciency in chloroplast electron transport processes and regeneration of ribulose‐1,5‐bisphosphate), and the initial slopes of the curve (an estimate of carboxylation eYciency). Caution is needed in producing these curves because a nonhomogeneous distribution of open stomata known as patchiness, as may occur under drought conditions, can lead to overestimation of calculated Ci with consequent problems interpreting the A:Ci relationship in some species. Wise et al. (1992), however, demonstrated that stomatal

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patchiness did not occur on water‐stressed field‐grown upland cotton or in drought‐acclimated growth chamber‐grown cotton. A homogeneous distribution of 14CO2 fixation in the leaves was also found for cotton grown at two levels of salinity (Plaut and Federman, 1991). These findings indicate that accurate estimates of the A:Ci relationship can be made in cotton under both well‐watered and water‐stressed conditions, by using proper care and methodology. Early estimates of a CO2 compensation point for cotton put it at 70 ml liter1 (Benedict et al., 1972). Krizek (1986), on the other hand, reported that the CO2 compensation point for cotton ranged from 60 to 120 ml liter1. Thomas et al. (1993) reported CO2 compensation points in the absence of nonphotorespiratory respiration of 44 ml liter1 when cotton was grown at two CO2 levels (350 and 650 ml liter1). Such discrepancies in the CO2 compensation values may be due to diVerences in temperature when measurements were taken, as Perry et al. (1983) showed that the rate of photorespiration is highly temperature dependent in cotton. Recent reports of A:Ci measurements suggest that variation in photosynthetic capacity within the chloroplast may exist among cotton varieties. Pettigrew and Turley (1998) found the CO2 compensation point diVered among six upland cotton genotypes in 1 year of a 2‐year study with ‘‘Dixie King,’’ a low CER genotype, also having a high CO2 compensation point. The CO2 compensation points in this study ranged in value from 62 to 69 ml liter1. The overall A:Ci curves for the six genotypes used in this study also diVered significantly in both years of the study. ‘‘Pee Dee 3’’ consistently produced a higher Amax, compared to the other genotypes. In a study of two short season cotton cultivars having common ancestry, Faver et al. (1996) found that the initial slope and Amax from A:Ci curves were greater for cultivar ‘‘TAMCOT HQ95’’ than ‘‘G&P 74 þ,’’ over water stress levels ranging from nonstressed to severe (leaf water potentials from 0.8 to 3.0 MPa). Their results suggest that the carboxylation eYciency (e.g., Rubisco activity) and the electron transport capacity or RuBP regeneration capacity in the chloroplast were greater for ‘‘TAMCOT HQ95’’ than ‘‘G&P 74þ.’’ In addition, boll retention and yield were higher for ‘‘TAMCOT HQ95’’ than ‘‘G&P 74þ’’ (Gerik et al., 1996). Thus, genetic improvements in photosynthetic capacity at the chloroplast level may be possible and lead to higher yielding cotton varieties, providing that this additional assimilate is partitioned into reproductive growth.

E.

13

C DISCRIMINATION

Water vapor and CO2 are the two primary gases exchanged by leaves with the atmosphere through the stomata. Plants lose H2O vapor through transpiration, take up CO2 via photosynthesis, and lose CO2 through respiration

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(both dark and photorespiration). During photosynthesis, CO2 uptake and H2O vapor loss occur simultaneously. The ratio of CO2 taken up to H2O vapor lost by the leaf is often referred to as WUE. Three methodologies are currently used to estimate plant WUE. First, the gas exchange method measures the instantaneous loss of H2O vapor and the CO2 taken up per unit leaf area. The second method involves the realized amount of dry matter produced by a plant per unit of water transpired. Finally, the ratio of 13C/12C isotopes present in the leaf dry mass relative to their natural abundance in the atmosphere (carbon isotope discrimination technique) estimates the intrinsic WUE of a leaf through the period of carbon accumulation. This method is based on Rubisco preferentially using CO2 with the lighter 12C isotope rather than the heavier 13C isotope and the theory that the level of discrimination decreases as Ci decreases, so plants with greater WUE may exhibit lower 13C discrimination in dry matter (Farquhar et al., 1989). Genotypic diVerences have been detected for WUE in cotton. As mentioned previously, lower gs combined with higher CER resulted in higher WUE, as determined through gas exchange, for super‐okra and okra leaf type isolines of MD 65–11 when compared to the normal leaf type isoline (Pettigrew et al., 1993b). Using the realized approach (e.g., dry matter production/water consumed) Quisenberry and McMichael (1991) found genotypic diVerences in WUE among greenhouse‐grown cotton genotypes. These diVerences were attributed to genetic diVerences in biomass production because genotypes used similar amounts of water when grown in pots. Furthermore, Cook and El‐Zik (1993) demonstrated that ‘‘Tamcot CD3H,’’ a relatively drought‐tolerant genotype, had a higher WUE than ‘‘Paymaster 303,’’ a drought susceptible genotype. Saranga et al. (1998) documented WUE diVerences among G. hirsutum cultivars, G. barbadense cultivars, and an interspecific F1 hybrid (G. hirsutum G. barbadense) and further correlated these WUE values with the carbon isotope ratio. Genotypic diVerences have been detected in the 13CO2 discrimination, and some of these diVerences have been utilized in breeding programs toward a goal of improving cotton water use eYciency. Gerik et al. (1995a,b) demonstrated carbon isotope discrimination as an indirect method to detect genotypic diVerences in WUE among 10 upland cotton genotypes. The upland variety ‘‘DPL Acala 90’’ was reported to have a significantly lower carbon isotope discrimination value than the Pima cotton varieties ‘‘S‐6’’ or ‘‘S‐7’’ (McDaniel, 1995). McDaniel (1995) also observed transgressive segregation (progeny with values both exceeding and less than their parental lines) for the carbon isotope discrimination trait across four generations of progeny from an interspecific cross between the upland and Pima cotton parents. Leidi et al. (1999) detected genotypic variation in 13CO2 discrimination

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among a diverse group of 27 G. hirsutum cultivars under dryland conditions. In addition, Stiller et al. (2005) were also able to document 13CO2 discrimination diVerences among G. hirsutum varieties. They were further able to calculate a broad sense heritability estimate of 0.68 for 13CO2 discrimination, similar to the estimates of 0.65 for net photosynthesis and 0.56 for yield. While Saranga et al. (2004) were able to identify 33 quantitative trait loci (QTL) for 13CO2 discrimination in two generations of progeny for a cross between G. barbadense and G. hirsutum, there was only incidental association between these QTL and productivity. This disconnect illustrates the lack of a direct relationship between WUE and yield. Carbon isotope discrimination among Pima cotton varieties was correlated with lint yield and heat resistance (Lu et al., 1996). They also found substantial genetic variation in the G. barbadense germplasm by using 11 uncultivated primitive accessions. Because carbon isotope discrimination was correlated with gs in commercial varieties, in a segregating population, and in a collection of primitive accessions, Lu et al. (1996) concluded that gs is the main source of variation for carbon isotope discrimination in Pima cottons.

III.

MANAGEMENT AND ENVIRONMENTAL EFFECTS

Genetic aspects of cotton photosynthesis can often be overshadowed or overwhelmed by environment or by management‐related alterations in the cotton crop canopy or plant photosynthetic capacity. While there may not be many things a producer can do to improve the crop’s potential photosynthesis capacity, there are ways the producer can provide for more optimum growing conditions to mitigate factors that have a negative impact on crop growth and photosynthesis.

A. PLANT GROWTH REGULATORS One of the more common methods producers use to manipulate growth of the cotton plant is through the application of plant growth regulators. Perhaps the most popular growth regulator in current usage is mepiquat chloride (1,1‐dimethylpiperidinium chloride), commonly known by some of its brand names such as PixÒ (BASF Corp., Parsippany, NJ), Mepichlor (MicroFlo Co., Mulberry, FL), or Mepex (GriYn Corp., Valdosta, GA). The eVects of mepiquat chloride on photosynthesis have been inconsistent, much as the eVects of mepiquat chloride on yield have proven to be inconsistent. Hodges et al. (1991) measured canopy gross photosynthesis under

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four temperature regimes following mepiquat chloride treatment and found gross photosynthesis increased on the second day at all temperatures except for the lowest temperature regime (20/10 C). They also reported mepiquat chloride increased night‐time respiration rates at all temperatures except 30 C. On the other hand, transpiration, daily gross carbon uptake, daily carbon loss, and daily net carbon gains of whole cotton plants grown under well‐watered or water‐deficient conditions were found to be inhibited by mepiquat chloride (Ferna´ndez et al., 1992). Mepiquat chloride also reduced the plant WUE under water‐stressed conditions. Reddy et al. (1995a) also found leaf CER was decreased about 26% compared to the untreated control when the concentrations of mepiquat chloride in the treated plants’ leaves ranged between 0.01 and 0.02 mg g1 dry weight. Another plant growth regulator under investigation for cotton production systems is PGR‐IVÒ (Microflo, Mulberry, FL), consisting primarily of gibberellic acid, indolebutyric acid, and a proprietary fermentation broth. Research utilizing PGR‐IV has reported increased photosynthesis from PGR‐IV treatments, when such diVerences were detected (Cadena and Cothren, 1995; Cadena et al., 1994; Guo et al., 1994; Oosterhuis and Zhao, 1993; Zhao and Oosterhuis, 1994, 1995). However, beneficial eVects of PGR‐IV on photosynthesis of upland cotton leaf seems to occur most often when the plants are under stress (drought, flooding, or nutrient). Leaves from water‐deficient plants treated with PGR‐IV photosynthesized at higher rates than nontreated plants (Cadena and Cothren, 1995; Guo et al., 1994; Zhao and Oosterhuis, 1994, 1995). The same seems to be true for cotton grown under flooded conditions (Zhao and Oosterhuis, 1994, 1995). Cadena et al. (1994) reported higher net carbon uptake of PGR‐IV‐treated cotton plants at 20 C but not at 30 C. PGR‐IV did not alter photosynthesis when plants were fertilized, but when not fertilized, PGR‐IV‐treated plants had greater photosynthesis than untreated plants (Cadena and Cothren, 1995). The eVects of other plant growth regulators on photosynthesis have been evaluated in cotton. Atonik (Asahi Chem. Mfg. Co., Osaka, Japan), composed of the sodium salts of ortho‐nitrophenol, para‐nitrophenol, and 5‐nitroguaiacol, was reported to increase photosynthesis (Guo and Oosterhuis, 1995), although some other plant growth regulators did not significantly aVect the photosynthetic rate (Guo et al., 1994). Ethephon [(2‐chloroethyl) phosphonic acid], marketed in cotton under the brand name Prep (Rhone‐ Poulenc Ag. Co., Research Triangle Park, NC), is used as a harvest aid to accelerate boll dehiscence prior to harvest. When ethephon was applied to induce early season square loss and theoretically stimulate the subsequent flowering, the treated plants exhibited lower CER, gs, and net assimilation rate than the control plants shortly after treatment (Pettigrew et al., 1993a).

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B. PLANT NUTRITION

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SOIL FERTILITY

The fertilizer regime under which cotton plants are grown can aVect the overall photosynthetic performance of the crop. Maintaining proper levels of soil fertility is a strategy producers can readily use to alleviate nutrient deficiencies that compromise photosynthesis and overall yield. The impact of all essential elements on the photosynthetic process is beyond the scope of this chapter, therefore we will concentrate on only the three major nutrient elements: nitrogen (N), phosphorus (P), and potassium (K). Because Radin and Mauney (1986) reviewed the physiological eVects of N nutrition on cotton growth, we will limit this discussion to N eVects on photosynthesis and closely related physiological traits. The predominant eVect of N stress on cotton photosynthesis is imposed by reductions in leaf area expansion rates (Radin and Parker, 1979b; Wullschleger and Oosterhuis, 1990), final leaf area (Radin and Parker, 1979a), and ultimately the crop’s leaf area index (LAI; Wullschleger and Oosterhuis, 1990). However, N deficiency does decrease leaf CER (Longstreth and Nobel, 1980; Radin and Ackerson, 1981) and net assimilation rates (Radin, 1983). Reddy et al. (1996) found that CER, gs, and Rubisco activity were all positively correlated with leaf N concentration in Pima cotton. Field‐grown cotton plants behaved diVerently, however, as Wullschleger and Oosterhuis (1990) failed to find any leaf CER diVerences between plants receiving adequate and deficient levels of N fertilization. Radin (1983) hypothesized that for cotton and other dicotyledonous plants, the eVect of N deficiency on leaf CER was secondary to that of the reduction in leaf area that it induces (i.e., it could decrease leaf area without eVecting CER on a leaf area basis). However, an N deficiency drastic enough to severely lower leaf N concentration would undoubtedly impair both leaf area expansion and leaf photosynthesis. In a related matter, stomata from leaves of N deficient‐cotton plants were prone to earlier closure under moisture deficit conditions than those from N suYcient plants (Radin and Parker, 1979b; Radin et al., 1984). Thus, under moisture deficit stress, N deficiency promotes water savings in cotton at the expense of photosynthesis and dry matter production. A deficiency in K can produce a similar eVect on photosynthesis as observed for N. Plants exhibited lower leaf CER under low K conditions (Bednarz et al., 1995a,b; Longstreth and Nobel, 1980; Pervez et al., 2004). Undoubtedly, this eVect is related to the close association between internal K levels and stomatal conductance, in that the mechanism of stomatal opening and closure is dependent on the flux of K into and out of the stomatal guard cells (Fischer, 1968; Fischer and Hsiao, 1968; Willmer and Pallas, 1972). Longstreth and Nobel (1980), however, concluded that the reduction in leaf CER due to K deficiency in cotton was primarily related to

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increased mesophyll resistance. Like N, the predominant eVect of K deficiency on the photosynthetic performance of a cotton crop is in the reduction of leaf area (Cassman et al., 1989; Mullins et al., 1994; Pettigrew and Meredith, 1997). Although not specifically documented in cotton, K deficiency reduced maximum leaf area expansion (final area) of individual leaves on soybean [Glycine max (L.) Merr.] plants (Huber, 1985). The translocation of the sugars produced via the photosynthetic process from the leaves to other plant parts may also be reduced by K deficiency (Ashley and Goodson, 1972). This reduction in the plant photosynthetic capacity (reductions in leaf area and CER per unit leaf area) is a principle cause of the yield reductions associated with K deficiency in cotton. A deficiency in P can also adversely aVect photosynthetic performance. While the total pool of phosphate (organic and inorganic) in the chloroplast is considered to be constant, inadequate triose phosphate utilization can cause the level of free phosphate in the chloroplast to decline and begin to limit photosynthesis (Sharkey, 1985). The manner in which low free phosphate concentrations or triose phosphate utilization limits photosynthesis is not clear, but Sharkey (1985) hypothesized that it was probably related to a reduction in the activation state of Rubisco. The inorganic phosphate level also regulates the transport of triose phosphate sugars out of the chloroplasts via the phosphate translocator, contributing to the utilization of the triose phosphates (Heber and Heldt, 1981). Phosphorus deficiency eVects on photosynthesis have been well documented in cotton. For cotton grown under ambient and elevated CO2 levels, there was a reduction of CER and gs associated with low P fertility levels imposed during growth (Ackerson, 1985; Barrett and GiVord, 1995; Longstreth and Nobel, 1980; Pettigrew et al., 1990a; Radin, 1984; Radin and Eidenbock, 1986). In addition to lower CER, Barrett and GiVord (1995) also reported lower relative growth rates for P‐deficient cotton. Although Barrett and GiVord (1995) reported lower rates of CO2 carboxylation under P deficiency, they hypothesized that photosynthesis was not limited by triose phosphate transport from the chloroplast because photosynthesis increased at all Ci greater than 500 ml liter1 for all the P treatments. This hypothesis contrasts with the data of Harley et al. (1992) suggesting the operation of a triose phosphate limitation in cotton grown at elevated CO2 and adequate P. The eVect of P deficiency on leaf starch concentration is not consistent. Radin and Eidenbock (1986) found increased predawn leaf starch levels in low P plants, but Barrett and GiVord (1995) reported lower leaf starch concentrations under phosphate deficiency. Although Ackerson (1985) reported increased leaf starch concentrations for low P treatments with drought stress‐acclimated cotton plants, the leaf starch levels of the nonacclimated control plants did not show a similar response to the P treatments when subjected to dehydration. Similar to the influences of N and K deficiencies, a P deficiency inhibits the rate of leaf expansion (Radin and Eidenbock, 1984).

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C. MOISTURE STRESS Moisture stress (particularly drought stress) has been extensively researched in many plant species, including cotton. Several reviews have outlined the diminished rate of net CO2 assimilation in water‐stressed plants (Ort and Boyer, 1985; Passioura, 1994; Sharkey, 1985). In this chapter, we will not go into extensive detail on the drought stress response in cotton but will merely highlight some of the more pertinent recent research. When the moisture deficits become severe enough, a reduction in the photosynthetic process occurs (Ackerson and Hebert, 1981; Ackerson et al., 1977; Ephrath et al., 1993; Plaut and Federman, 1991). Under moderate moisture stress conditions, where the soil is able to maintain some semblance of the plants hydraulic status, Pettigrew (2004a) documented a higher photosynthetic rate for drought‐ stressed plants during the morning before declining in the afternoon to a point lower than the irrigated control plants as the hydraulic status deteriorated. Reduced leaf size caused by the moisture stress was speculated to have led to more concentration of the photosynthetic machinery per unit leaf area which was able to sustain a greater photosynthetic rate during the morning while the plants hydraulic status was still acceptable. The mechanism driving this moisture deficit stress‐induced photosynthesis reduction is not completely clear, since both stomatal and nonstomatal restrictions have been implicated. In greenhouse or growth‐chamber studies, both gs and CER were reduced as water deficits decreased the leaf water potential (c) (Ackerson and Hebert, 1981; Plaut and Federman, 1991; Radin and Ackerson, 1981). Nonstomatal factors have also been identified as contributing to the photosynthetic reduction in these greenhouse‐grown plants under drought stress (Ackerson and Hebert, 1981; Plaut and Federman, 1991; Wise et al., 1992). Repeated exposure to drought stress in pot‐grown cotton plants can lead to a photosynthetic acclimation which causes a delay in the photosynthetic reduction with subsequent water deficits (Ackerson and Hebert, 1981; Plaut and Federman, 1991). In field‐grown cotton, the drought stress eVect on gs is not consistent. Ackerson et al. (1977) reported drought stress had little eVect on the gs of field‐ grown cotton, but CER was substantially reduced. Their data suggested that nonstomatal factors were primarily responsible for the photosynthetic declines seen in the plants. On the other hand, Ephrath et al. (1993) found both CER and gs were reduced by water deficit. Because gs declined and Ci increased under drought stress conditions, they concluded that nonstomatal components are the main limiting factors to CER for field‐grown cotton plants under moisture deficit stress. The fact that stomatal and nonstomatal processes operate concurrently may have contributed to the confusion on the major impact of these factors on photosynthesis of water‐stressed cotton. Using gas exchange methods, Faver et al. (1996) found that both factors were important in reducing

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photosynthesis when cotton was water stressed. Under mild water stress, nonstomatal factors were most important in reducing photosynthesis. As leaf water potentials declined from 0.8 to 1.5 MPa, small reductions in Ci and A were observed (<12%), with substantial reductions in gs (45%). As water stress became severe and leaf water potentials declined below 1.5 MPa, reductions in A and Ci paralleled reductions in gs. Thus, nonstomatal factors appeared to be more important in reducing photosynthesis when water stress was mild, while gs was more important when stress levels became severe. Mechanisms for nonstomatal reduction in photosynthesis by water stress are not known. However, A:Ci curves from leaves of water‐stressed cotton revealed substantial reduction in the A:Ci initial slope and Amax with water stress (Faver et al., 1996). Because the A:Ci initial slope is correlated with Rubisco activity and the Amax from the A:Ci curve is correlated with the electron transport capacity for RuBP regeneration (Farquhar et al., 1980), water stress appears to reduce both carboxylation and electron transport processes in the chloroplast. Reductions observed in the A:Ci initial slope (Rubisco activity) and Amax (RuBP regeneration) were linear with increasing water stress–leaf water potentials ranging from 0.8 to 1.5 MPa (Faver et al., 1996). These data support the Hutmacher and Krieg (1982) conclusion that nonstomatal factors have a major influence on the assimilation capacity of water‐stressed cotton. Radin (1992) addressed the inconsistency of the gs response between field and controlled environment chambers and attributed it to temperature eVects of the abscisic acid (ABA) concentration in the leaves. Raschke (1975) had demonstrated previously that the eVect of ABA on the stomatal aperture was dependent on internal CO2 and vice versa. High temperatures decrease the accumulation of ABA in cotton leaves during water stress (Radin et al., 1982). These two facts were brought together in the theory that high temperatures in the field lowers the ABA concentration in cotton leaves, which in turn causes the stomata to be less responsive to internal CO2 and uncoupled from mesophyll photosynthetic components. While not proving cause and eVect or accounting for all other environmental and developmental factors, this theory does appear to explain the majority of the diVerences in the gs response seen between field and controlled environment studies. While prolonged exposure to excessive moisture of flooding conditions undoubtedly has a negative impact on plant growth and probably photosynthesis, we are unaware of any research in cotton that directly addresses the photosynthetic response. Problems associated with having suYcient water during critical periods of growth and development of the cotton crop have led to questions about, and the desire to improve the eYciency with which cotton uses available water.

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The genetic variability of various components of WUE was discussed earlier in this chapter. Using the realized approach (dry matter production/water consumed), Cook and El‐Zik (1993) demonstrated that water stress resulted in 45% reduction in WUE, compared to the nonstressed control. This finding was probably related to stress‐induced nonstomatal inhibition of photosynthesis or reduced leaf area expansion. There is probably an optimal range of soil water availability for cotton, above and below which the WUE of the crop begins to decline. Hypothetically, a goal for producers might be to maintain the WUE of their crop at a maximal level for as long as possible. This goal assumes that any increased growth will be translated into increased yield, which is not necessarily an automatic consequence. However, the consequence of rising atmospheric CO2 levels that have occurred in recent years, and are expected to continue into the foreseeable future are expected to influence WUE due to the eVect of CO2 levels on gs and hence, transpiration. Reddy et al. (1995b) found that cotton grown at 2 ambient CO2 concentrations (700 ml liter1) had greater WUE than cotton grown at ambient levels (350 ml liter1). This response was primarily due to increased photosynthesis at the higher CO2 but also, to a lesser degree, to reduced transpiration in higher CO2.

D. TEMPERATURE Due to the large geographic region encompassed by the US cotton production belt and the normal progression of temperatures across the growing season at any particular site, growth temperature extremes on either side of the optimal range can be experienced by a US cotton crop. These can range from the excessively high temperatures of the Arizona and California deserts to periods of exceedingly cool night temperatures for the Texas high plains and northern fringes of the US cotton production belt (Missouri Bootheel, northern Tennessee, northern North Carolina, and Virginia). Like all biochemical reactions, there is an optimum temperature range to achieve sustained maximum photosynthetic metabolic activity. The temperature optimums for general growth and yield in cotton are reported to range from 23.5 to 32 C with the best temperature being 28 C (Burke et al., 1988). Focusing specifically on photosynthesis, Perry et al. (1983) showed that gross photosynthesis in upland cotton had a curvilinear response to the air temperature during the photosynthesis measurements between 22 and 40 C, with a peak at about 32 C. They also showed a photorespiratory response across the same temperature range that increased almost linearly, so consequently net photosynthesis exhibited a linear decline as the temperature was elevated above 22 C.

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Considerable research in recent years has been directed toward identifying some of the components of the overall photosynthetic process that contribute to the reduced performance of photosynthesis at suboptimal temperatures. Warner and Burke (1993) showed that cotton grown at a 28/20 C (day/night) temperature regime had increased predawn leaf starch concentrations and decreased chlorophyll Fv/Fm fluorescence ratios, compared to control plants grown at a constant 28 C. Because they were unable to separate the increased starch concentrations from the decreased chlorophyll fluorescence among the diVerent treatments, they concluded that reduced starch metabolism under the cool nights led to the elevated predawn starch levels, which inhibited PS II activity as indicated by chlorophyll Fv/Fm. Cotton plants grown under low light (450–540 mmol m2 s1) and a normal temperature regime 28/22 C (day/night) but later exposed to moderate constant chilling 15 C, exhibited a 40% reduction in CER, an 88% decline in gs, a 33% reduction in Ci, and a chlorophyll fluorescence Fv/Fm ratio that was 29% lower than prechilling values (Perera et al., 1995). Initial Rubisco activity remained relatively stable during the early stages of the chilling treatment, before beginning to decline during the later stages of chilling. Initial activity of stromal fructose 1,6‐bisphosphatase increased during the early part of the chilling period, before becoming stable. Leaf concentrations of glucose 6‐phosphate, fructose 6‐phosphate, starch, sucrose, glucose, and fructose for these plants all increased during the chilling treatment. Activities of leaf ADP‐glucose pyrophosphorylase (for starch synthesis) and UDP‐glucose pyrophosphorylase (for sucrose synthesis), and cytosolic fructose 1,6‐bisphosphatase (for sucrose synthesis) increased during chilling, if only slightly. Sucrose phosphate synthase, on the other hand, decreased during the course of the chilling treatment. Hexokinase activity changed little during the chilling. Perera et al. (1995) interpreted these data to mean that inhibition of CER during the early stages of the chilling was primarily due to photoinhibition and to a lesser degree to a gs limitation. They suggested that CER inhibition during the later stages of chilling was further exacerbated by a phosphate limitation to photosynthesis. Although the CER, dry matter production, gs, O2 evolution, and Fv/Fm were depressed when cotton was grown at lower day time temperatures across the range 20–34 C, the temperature at which the decline started varied depending on the light levels during growth (Ko¨niger and Winter, 1993a; Winter and Ko¨niger, 1991). Leaf chlorophyll and carotenoid concentrations were reduced in the low‐temperature‐grown plants, but because the xanthophyll concentration remained stable over this range, its proportion of the carotenoids was increased at the lower temperatures (Ko¨niger and Winter, 1991). Maintaining a constant day temperature of 34 C while reducing the night temperature from 25 to 10 C caused a reduction in dry

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matter production and leaf area but did not alter O2 evolution, quantum yield, and Fv/Fm (Ko¨niger and Winter, 1993b). Keeping the root temperatures at 25 C while the night temperatures were reduced alleviated some of the depression in dry matter production and leaf area development, but did not aVect O2 evolution. Perera et al. (1995) also showed that maintaining the roots warm while reducing the air temperatures the cotton plants were exposed to had no eVect on CER and gs. Much like the deleterious eVect of cool temperatures, photosynthesis is also negatively impacted when temperatures become excessively high (Perry et al., 1983). Increased photorespiration (Perry et al., 1983) and dark respiration (Reddy et al., 1991) certainly contribute to the reduction in photosynthesis as temperatures increase. Although eVects of high‐temperature stress on photosynthetic components have not been adequately addressed in cotton, work in other species has shown that elements of both the dark and light reactions of photosynthesis can be damaged by excessively high temperatures. PS II is more sensitive to heat damage than PS I, while the noncyclic photophosphorylation is more labile under conditions of high temperature than cyclic photophosphorylation. Stromal components are generally more stable under high‐temperature stress than thylakoid components. However, photosynthesis above 40 C can be limited by the activation state of Rubisco (Paulsen, 1994). Recent work out of the USDA‐ARS Western Cotton Research Laboratory in Phoenix, AZ indicated that a reduction in the activation state of Rubisco can be a limiting factor to photosynthesis in cotton under heat‐ stress conditions. The actual level of Rubisco activity is due to a balance between the rate of Rubisco deactivation and reactivation by Rubisco activase. This decrease in activity as the temperature increases is because of both a faster Rubisco deactivation rate and a slower reactivation rate by Rubisco activase (Crafts‐Brandner and Law, 2000; Crafts‐Brandner and Salvucci, 2000, 2004; Law and Crafts‐Brandner, 1999; Salvucci and Crafts‐Brandner, 2004). In a counterargument to the Rubisco deactivation concept, recent work by another group failed to document any photosynthetic limitation caused in field‐grown Pima cotton by lack of Rubisco activity (Wise et al., 2004). They felt that the true photosynthetic limitation to Pima cotton experiencing heat stress was actually in the thylakoid reactions, with the corresponding Rubisco activity regulated to not greatly exceed the level of RuBP regenerated by the photosynthetic electron transport. Schrader et al. (2004) demonstrated that heat stress leads to leaky thylakoid membranes and increased PS I activity at the expense of stromal reductants. Developing genotypes that are able to grow and function more successfully under temperature extremes encountered during a growing season (without

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sacrificing optimal growth under normal conditions) should probably be added to the priority list of cotton breeding goals. However, when evaluating four upland cotton cultivars diVering in leaf shape and leaf pubescence, Bednarz and van Iersel (2001) were not able to detect variety diVerences in net photosynthesis or dark respiration as the temperature was increased from 6 to 34 C in 4 C increments. A phenomenon that has been documented in many plant species, that in some quarters has been at least partly attributed to high‐temperature stress, is the afternoon decline in photosynthesis compared to morning rates, even though comparable sunlight levels are encountered. A handful of studies have documented lower afternoon photosynthetic rates in both upland (Pettigrew, 2004a; Pettigrew and Turley, 1998; Pettigrew et al., 1990b) and Pima (Cornish et al., 1991) cotton. A number of processes could cause or contribute to this phenomenon. These include photoinhibition after exposure to the intense solar radiation experienced at solar noon (Powles, 1984), end‐product inhibition of the photosynthetic process due to buildup of large carbohydrate levels during the afternoon hours (Nafziger and Koller, 1976; Peet and Kramer, 1980), stomatal closure due to increasing H2O vapor pressure deficits in the afternoon hours (Bunce, 1982, 1983; Farquhar et al., 1980; Pettigrew et al., 1990b), nonstomatal photosynthetic inhibition due to transient, localized water stress caused by high transpiration demand in the afternoon (Sharkey, 1984), and high‐temperature stress (Baldocchi et al., 1981; Perry et al., 1983). Baker and Ort (1992) described situations where the eYciency of light utilization by the photosynthesis process decreased with increasing light intensity. This ‘‘downregulation of photosynthesis’’ is thought to be regulatory in nature, not caused by damage to the photosynthetic apparatus, and it is believed to redirect absorbed photons away from reaction centers and dissipate them as heat instead of allowing them to damage the photosynthetic apparatus. This intrinsic regulatory reduction of photosynthetic eYciency can often be mistakenly diagnosed as photoinhibition. Photoinhibition often involves damage to the PSII reaction center proteins and is not quickly reversed, while a regulatory reduction of the photosynthetic eYciency causes no damage to the photosynthetic system and is generally reversible within 1 h of being placed in darkness (Baker and Ort, 1992). Because of the complexity of factors and interactions involved in the afternoon photosynthetic inhibition, it is next to impossible to predict when this phenomenon will begin and to what extent it will depress leaf photosynthetic rates. When conducting photosynthetic studies, the possibility that lower rates could be encountered in the afternoon must to be taken into account, just like any other plant or environmental factor that can alter photosynthesis. Improving the overall photosynthesis during the afternoon hours (through genetics, production techniques, or whatever means possible) would seem to be an area that is ripe for physiological research. Any improvements in this area may eventually lead to potential yield increases.

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IV. CONCLUSIONS Considerable progress has been made in understanding photosynthesis and carbon metabolism in plants in general and in cotton specifically. However, our understanding of the relationship between instantaneous gas exchange measurements of photosynthesis and plant growth and development remains quite rudimentary. As with pest management, the tools of genetic engineering oVer the promise of potentially improving and furthering the understanding of both photosynthesis and reproductive sink capacity. How or whether the introduction of foreign or manufactured DNA into the plant genome will aVect nontarget native physiological processes is not clear. At the present time, there is no control over where the foreign DNA is inserted into the plant genome. On the other hand, our knowledge of rate‐limiting enzymes can possibly be improved by molecular genetic techniques which might not have been achieved through traditional methods. With such a large potential payoV, yet with so many unanswered questions but also with so many new tools to address these problems, this area of physiology in cotton is poised for important breakthroughs.

V. SUMMARY Photosynthesis is the process where plants use energy from sunlight to convert CO2, acquired from the air, water, and soil, into food for growth. It is one of the most essential processes in nature because it not only supports plant growth but it also releases oxygen into the atmosphere as a result of the process. This chapter complies information from numerous journal references and other sources to present the current state of the art and understanding of cotton leaf photosynthesis and carbon utilization. Cotton geneticists can make use of this information to implement breeding strategies with goal of improving cotton photosynthesis as a component toward yield improvement. Crop physiologists and agronomists can use this information in designing cotton production practices that enhance photosynthetic and yield performance by minimizing exposure to detrimental environmental influences. This chapter can also fortify the knowledge base and background that consultants and extension specialists draw on to advise their producer customers and clients.

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THE IMPACTS OF GRAZING ANIMALS ON THE QUALITY OF SOILS, VEGETATION, AND SURFACE WATERS IN INTENSIVELY MANAGED GRASSLANDS G. S. Bilotta,1,2 R. E. Brazier1 and P. M. Haygarth2 1

Department of Geography, University of Exeter, Exeter, Devon EX4 4RJ, United Kingdom 2 Cross Institute Programme for Sustainable Soil Function (SoilCIP), Institute of Grassland and Environmental Research (IGER), North Wyke Research Station, Devon EX20 2SB, United Kingdom

I. Introduction II. Impact of Treading by Grazing Animals on Grassland Soils III. Factors Influencing the Amount and Form of Soil Structural Alteration A. Animal Species and Age B. Stocking Density C. Soil Moisture D. Vegetation Cover IV. Forms of Soil Structural Alteration Resulting from Treading by Grazing Animals A. Soil Compaction B. Soil Pugging C. Soil Poaching V. The Impacts of Soil Structural Alteration by Grazing Animals A. Treading and Soil Physical Properties B. Treading and Soil Hydrology C. Treading and Vegetation Growth D. Treading and Soil Fauna VI. The Impact of Defoliation by Grazing Animals on Grassland Vegetation A. Animal Species and Age B. Stocking Density C. Vegetation Response VII. Impact of Excretion by Grazing Animals on Vegetation, Soils, and Surface Waters in Intensively Managed Grasslands A. Livestock Wastes as a Source of Nutrients B. Livestock Wastes as a Source of Pathogens

237 Advances in Agronomy, Volume 94 Copyright 2007, Elsevier Inc. All rights reserved. 0065-2113/07 $35.00 DOI: 10.1016/S0065-2113(06)94006-1

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G. S. BILOTTA ET AL. VIII. Impacts of Grazing Animals on the Water Quality of Surface Waters in Intensively Managed Grasslands A. Soil Erosion and Sedimentation Problems B. Eutrophication C. Pathogenic Contamination IX. Environmental Degradation by Grazing Animals: Recovery and Remediation A. Natural Recovery of Soil Physical Condition Following Treading Damage B. Mitigation and Damage Reduction Methods X. Future Research Acknowledgments References

This chapter provides a comprehensive review of the literature relating to the impacts of grazing animals on the quality of soils, vegetation, and surface waters. It focuses on intensively managed grasslands where there is the greatest potential for these impacts to be observed. The chapter indicates that while well‐managed grazing can be beneficial to the environment, intensively managed grazing can actually lead to the degradation of both the soil and vegetation of grassland environments. The various causes, forms, and consequences of this degradation are discussed in detail, and gaps in the knowledge are identified. The chapter highlights the need for recognition and quantification of the relationships between the on‐site impacts of grazing animals (i.e., changes in soil properties and vegetation cover) and the oV‐site impacts of grazing animals (i.e., the impact of these changes on hydrology and water quality in surface waters), as these relationships have, in the past, only been alluded to by authors. However, there exists relatively little research evidence to support and quantify these relationships, thus herein we describe data required to address the lack of understanding of the role of grazing animals on grasslands. Finally, the last section of this chapter considers the land management and remediation options available for the reduction of the impacts of intensive livestock farming. # 2007, Elsevier Inc.

I. INTRODUCTION Grasslands cover a large portion of the temperate landmass, including significant areas of Europe, North America, New Zealand, and Australia. In Western Europe, for example, grasslands occupy almost 40% of the agricultural area, although this proportion is even higher in some of the countries within Europe (e.g., Austria, 57%; Ireland, 76%; Switzerland, 72%; United Kingdom, 65%) (Peeters, 2004). Much of this grassland is grazed by livestock, providing its people with meat and dairy products, employment, and a source

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of income which in some regions is central to the local economy (Carroll et al., 2004; Reynolds and Frame, 2005). While well‐managed grazing can be beneficial to the environment, enhancing nutrient cycling and promoting floral and faunal biodiversity (Isselstein et al., 2005; Pykala, 2000; Rook and Tallowin, 2003), it is now recognized that intensively managed grazing can lead to environmental degradation (Evans, 1997, 1998; Heathwaite et al., 1990; Kellett, 1978; Mulholland and Fullen, 1991; Patto et al., 1978; Trimble and Mendel, 1995). Intensively managed grasslands tend to be located in lowland areas and are characterized by high stocking densities (i.e., high number of animals per unit area) and high inputs of chemical (e.g., fertilizers, pesticides, imported animal feeds) and energy (e.g., farm machinery, tractors) resources. These practices are designed to maximize agricultural productivity and are fairly widespread in temperate regions, with 29% of the total land area in England in Wales managed in this manner (Defra, 2005). However, intensive farming practices have been associated with changes in the percentage cover and biodiversity of grassland vegetation and alterations of the condition of the grassland soil. Together, these changes have been linked to a modified hydrological behavior of pastures and, ultimately, the deterioration of water quality in surface waters within these environments (Kurz et al., 2006; McDowell et al., 2003; Monaghan et al., 2005). The environmental degradation induced by grazing animals is a consequence of several key activities which livestock carry out, including defoliation, treading, and excretion. First, excessive defoliation by grazing animals and damage to plant tissues as a result of direct (e.g., crushing, bruising, shearing) and indirect (e.g., changes to the rhizosphere as a result of compaction, pugging, and poaching) treading eVects can reduce both the biodiversity of the pasture and the percentage cover of the vegetation (Matches, 1992). This can lead to a decline in faunal biodiversity and pasture productivity, and may eventually produce bare patches within the pasture where the soil surface is exposed to erosive agents (Edmond, 1958; Evans, 1998; Matches, 1992). Second, the treading action of livestock hooves on the soil surface, particularly on wet and saturated soils, can cause structural deformation of the soil such as compaction, pugging, and poaching, which can reduce the porosity of the soil and increase the bulk density of the soil (Climo and Richardson, 1984; Di et al., 2001; Drewry and Paton, 2005). This in turn can decrease the infiltration capacity (Mulholland and Fullen, 1991) and hydraulic conductivity of the soil (Greenwood et al., 1997; Willatt and Pullar, 1983), and can therefore promote surface runoV generation (Di et al., 2001; Heathwaite et al., 1990). If this occurs over large areas of a catchment, it can alter the hydrology of whole rivers (Carroll et al., 2004; Harrod and Theurer, 2002). Indeed, there is anecdotal evidence to suggest that the response of rivers to rainfall events is becoming more intense because of poor agricultural soil management practices (Reynolds et al., 2002),

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and a report by the UK Environment Agency (2002) estimated that the damage to homes, commercial property, and agricultural land resulting from the poor soil structure caused by intensive agriculture costs the UK approximately £115 million per annum (in 2000 prices). Furthermore, this enhanced runoV can potentially mobilize large amounts of sediment and colloidal material (including soil, plant, and livestock fecal matter) from the damaged and exposed soil surface of the grassland, and deliver this matter into surface waters where it could contribute to sedimentation problems (Harrod and Theurer, 2002; Walling et al., 2003), eutrophication (Haygarth and Jarvis, 1999; Heathwaite and Johnes, 1996), and pathogenic contamination (Chadwick and Chen, 2002; Oliver et al., 2005b). The deterioration of water quality induced by intensive grazing is of particular concern to those involved in meeting the demands of environmental legislation such as the EU Water Framework Directive (Neal and Jarvie, 2005). The environmental degradation of grasslands induced by grazing animals has become more prominent over the last few decades and has been associated with the intensification of agricultural production and the gradual decline in the area under grassland, while livestock numbers have been maintained or have increased, resulting in higher stocking densities on the remaining grassland area (Evans, 1997; Heathwaite et al., 1990). This chapter provides a review of the literature on the eVects of grazing animals on the quality of soils, vegetation, and surface waters in intensively managed, temperate grasslands, synthesizing the key findings, and identifying those areas where further research is required.

II. IMPACT OF TREADING BY GRAZING ANIMALS ON GRASSLAND SOILS Soil is composed of inorganic and organic primary particles, the size distribution of which determines the texture of the soil. Primary particles may be bound together to form aggregates, the size and arrangement of which determine the volume and configuration of spaces and pores within the soil and constitute collectively, the structure of the soil. The resistance of this soil structure to an imposed force (i.e., the shear strength of the soil) results from internal friction and interlocking of primary particles and aggregates, supplemented by inter‐particulate cohesion and bonding (Patto et al., 1978). When a load imposed on the soil (i.e., a shear stress) is greater than the load‐bearing capacity of the soil, it will lead to a modification of the structural configuration (i.e., soil deformation). Grazing animals can exert a large amount of force on the soil surface due to their large weight and relatively small hoof area. The amount of pressure exerted on the soil is dependent on the species and age of the grazing animal. The amount and form of soil structural alteration

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(i.e., deformation) which occurs as a result of this force is primarily determined by the stocking density, soil moisture content, soil texture, and the presence/ absence of a protective vegetation cover. This section of the chapter first examines the key factors that influence the amount and form of soil deformation, and then moves on to discuss the individual forms of soil deformation. This is then followed by an evaluation of the likely implications that these changes may have on the wider environment.

III. FACTORS INFLUENCING THE AMOUNT AND FORM OF SOIL STRUCTURAL ALTERATION A. ANIMAL SPECIES

AND

AGE

The force imposed on the soil by a grazing animal is a function of the weight of the animal and the area of contact between the animal hooves and the soil surface (Patto et al., 1978). Clearly, this will vary depending on the species and age of the animal, with cattle exerting the greatest forces onto the soil. For example, an adult cow weighs 350–600 kg (Abdel‐Magid et al., 1987; Mulholland and Fullen, 1991; Scholefield and Hall, 1986; Wind and Schothorst, 1964). When a cow is static, this mass will be distributed over four hooves, each with an area of around 60–90 cm2 (Frame, 1971), depending on cow breed and age (Scholefield and Hall, 1986). This creates static pressures of around 200 kPa (Di et al., 2001; Willatt and Pullar, 1983). These forces may be significantly increased when the animal is walking and has only two or three hooves in contact with the ground at any one time (Di et al., 2001; Trimble and Mendel, 1995; Willatt and Pullar, 1983), leading to forces of up to 400 kPa (Climo and Richardson, 1984). The forces may increase again if the hoof is not placed onto a flat surface (Di et al., 2001; Willatt and Pullar, 1983). For sheep, the body mass and resultant hoof forces are much lower than those for cattle, ranging from 50 to 80 kPa while static (Noble and Tongway, 1986; Willatt and Pullar, 1983), and up to 200 kPa when moving (Willatt and Pullar, 1983). Nevertheless, both the forces from sheep and cattle exceed those recorded from tractors which are known to cause compaction, yet exert forces ranging from just 30 to 150 kPa (Soane, 1970; Soehne, 1958). Moreover, the soil compaction caused by grazing animals is likely to be more widespread within paddocks than that caused by vehicle tracks (Drewry, 2006). The force imposed by an animal hoof can be divided into two components; a normal component acting vertically downward onto the soil and a tangential component acting horizontally to the soil surface (Patto et al., 1978). It is this normal component of force, which occurs as the hoof is placed down onto the

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soil surface, which is responsible for soil compaction, pugging, and poaching (Greenwood and McKenzie, 2001). The tangential component tends to occur as the hoof is lifted for a subsequent step and causes a shearing of the soil which can smear the soil and tear vegetation (Alexandrou and Earl, 1997).

B.

STOCKING DENSITY

Several researchers have found that the amount of soil structural alteration induced by grazing animals increases with stocking density (the number of animals per unit area) (Bryant et al., 1972; Langlands and Bennett, 1973; Mulholland and Fullen, 1991; Willatt and Pullar, 1983). This has been attributed to the fact that as the stocking density increases, the frequency at which any given point in the paddock/pasture is visited by a grazing animal also increases. Each time a point is revisited, it leads to further breakdown of soil structure and water‐stable aggregates, making the soil more susceptible to further alteration (Patto et al., 1978; Wind and Schothorst, 1964). Kellett (1978) proposed that increased soil structural alteration with increasing stocking density is also a result of the lower vegetation/protective cover available at higher stocking densities, where there are greater rates of defoliation. In contrast to the above findings, a 30‐year study by Greenwood et al. (1997) found no evidence of increased structural alteration with increased stocking density. Greenwood et al. (1997) propose that this may be because the eVects of grazing animals are cumulative and therefore tend to reach a common state over the long‐term. The magnitude of the relationship between soil damage and stocking density reported in the literature varies significantly between studies. This may be due to diVerences in the soil health indicator measured (e.g., bulk density, infiltration capacity, porosity, or macroporosity), diVerences in the methods by which grazing intensities were simulated/produced, diVerences in soil type, topography, and climate, diVerences in stocking management, and diVerences in the duration of experiments and observations. Trimble and Mendel’s (1995) review of the literature revealed that there are strong diVerences in the methodologies used in livestock impact studies. For example, some studies are carried out on natural watersheds using natural rainfall events, while others utilize small plots and/or flumes and artificial rainfall (Trimble and Mendel, 1995). In one case, a storm equivalent to the 150‐year return period was required to produce overland flow from poached land—reducing the confidence in the results (Trimble and Mendel, 1995). Perhaps more importantly, few simulation studies calibrated their means of simulating the eVect of livestock with the real animals (Trimble and Mendel, 1995), with the notable exceptions of workers such as Scholefield and Hall (1986). Furthermore, when researchers used outside plots/catchments, there was little consideration of the influence that

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prior land treatment, such as history of grazing, could have on the results (Trimble and Mendel, 1995). There is a paucity of knowledge about such lag eVects, but they may be significant and may persist for decades (Greenwood et al., 1997). Trimble and Lund (1982) and Trimble (1988, 1990) suggest that the eVects of land abuse and land recovery may take anything from several years to several decades to manifest themselves. In the existing literature describing the eVects of grazing intensity/stocking density on grasslands, it is clear that there are no universal definitions of treatments in terms of stocking rates, duration, and seasonality (Trimble and Mendel, 1995). This can create diYculties when attempting to make cross‐ study comparisons (Trimble and Mendel, 1995). One standardized measure of stocking density is Livestock Units (LSU) per hectare (Carroll et al., 2004). For example, using this definition, a dairy cow is equivalent to 1 LSU, whereas a sheep is regarded as 0.15 LSU (Carroll et al., 2004). However, Evans (1998) argues that when stocking density is given in terms of LSU, the comparison is made even more diYcult, because it is not known how many animals are grazing an area, for the proportions of the diVerent kinds of animals may vary but give the same LSU intensity. Evans (1998) suggests that if grazing intensity is given in terms of the animal, generally cattle or sheep, intensities can be compared between localities. There is also an issue over the style of livestock management, that is continuous stocking versus rotational stocking. Some authors include the length of grazing time per stocking density, others just mention the stocking density without a reference to the grazing period. This can lead to discrepancies in findings from each study. The majority of the existing research into the impact of livestock and stocking densities on grasslands has taken place in New Zealand, Australia, and America (Carroll et al., 2004). At present, there are very little quantitative data available for other countries such as the United Kingdom (Carroll et al., 2004; Trimble and Mendel, 1995). While the data from outside the country of interest can be useful, it must be used with caution as there may be issues of transferability due to diVerences in climate, soil type, vegetation, and grazing management style (Trimble and Mendel, 1995). Since it is estimated that grasslands cover around two‐thirds of the UK land area (Carroll et al., 2004; Reynolds and Frame, 2005; Waters, 1994), the impact of stocking density is an important area that requires further research.

C. SOIL MOISTURE Several authors have reported that soil damage induced by grazing animals is worsened by increasing the moisture content of the soil (Climo and Richardson, 1984; Mulholland and Fullen, 1991; Scholefield and Hall, 1986; Wind and Schothorst, 1964). As mentioned previously, the resistance of a soil

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structure to an imposed force results from internal friction and interlocking of primary particles and aggregates, supplemented by interparticular cohesion and bonding (Patto et al., 1978). Under soil conditions of low suction or positive water pressure, the eVective contact between soil particles (a source of internal friction) is reduced and particles can move more easily along failure planes, reducing soil strength and resistance to structural alteration (Patto et al., 1978). The soil moisture content will also determine the dominant form of soil structural alteration. As a general rule, soil compaction tends to dominate at low to medium soil moisture contents, followed by pugging at the medium to high moisture contents, and poaching on saturated soils (Mulholland and Fullen, 1991). Scholefield and Hall (1985) suggest that although there is a general assumption that the extent of soil deformation is primarily determined simply by the soil water content, their research supports a more complex model. Scholefield and Hall (1985) found that over a wide range of water contents, deformation due to treading was independent of water content. These results and those of other studies (Mullen et al., 1974) support a model for poaching alluded by Mullins and Fraser (1980) in which soil strength declines progressively during repeated treading in the presence of free water (i.e., water on and within the soil surface, not held in soil pores). Therefore, with this model, the amount of deformation will be determined by the rate of loss of soil strength during treading in wet weather. This rate will be determined by the rate at which water can be incorporated into the soil (which in turn is determined by soil texture and the amount of existing deformation), and also by the sensitivity of interparticular bonds (which in turn is determined by soil mineralogy and organic content) to mechanical disturbance, neither of which may be predicted by a single measurement of the initial state of the soil. This highlights the importance of soil texture and mineralogy, as well as the number of treading instances during wet weather, not simply initial soil moisture content.

D. VEGETATION COVER The protective role of plant cover with respect to damage by animal hooves on the soil has long been recognized (Climo and Richardson, 1984; Kellett, 1978; Scholefield and Hall, 1986). The protection to the soil oVered by plants is derived in several ways; first, the above‐ground plant matter provides a direct physical boundary between the hooves and the soil (O’Connor, 1956). Second, the below‐ground plant matter (roots and stolons) in the soil acts to increase the shear strength of the soil and its load‐bearing capacity (Patto et al., 1978). Third, plants provide protection for the soil indirectly through the decomposition of plant residues which bind with the mineral component of the soil and together with other agents, such as calcium/magnesium carbonates, iron/aluminum

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oxides, and silicates, give rise to water stable aggregates which are more resistant to deformation (Patto et al., 1978; Taylor and Ashcroft, 1972; Tisdale and Oades, 1982). The degree of protection oVered by vegetation depends on the quality and quantity of vegetation (Climo and Richardson, 1984). Established dense turf mats, which often form under low fertility conditions, provide good physical protection to the soil (Sears, 1956). Higher producing pastures, where fertilizer use is more common and the sward is dominated by species such as perennial ryegrass (Lolium perenne) and white clover (Trifolium repens), tend to be more open and allow direct hoof/soil contact and so oVer a lower degree of protection than a dense nutrient‐poor grassland sward (Climo and Richardson, 1984). It is well known that grassland species vary in their ability to survive trampling and, therefore, in the degree of protection that they provide for the soil surface (Patto et al., 1978). For example, a study in New Zealand by Edmond (1962) found perennial ryegrass to be far more tolerant of treading than white clover.

IV. FORMS OF SOIL STRUCTURAL ALTERATION RESULTING FROM TREADING BY GRAZING ANIMALS There are three main forms of structural alteration associated with grazing animals, these are: compaction, pugging, and poaching. These forms have sometimes been bulked together by authors and referred to as ‘‘poaching’’ (Kellett, 1978). However, in this chapter, the impacts of grazing animals are treated as individual processes because they operate under diVerent conditions and can have diVerent eVects.

A.

SOIL COMPACTION

Soil compaction has traditionally been defined as the compression of an unsaturated soil body resulting in a reduction of the fractional air volume (Hillel, 1980). The potential for grazing animals to cause soil compaction was first noted by authors such as Tanner and Mamaril (1959) and Federer et al. (1961). Soil compaction occurs when the load of a grazing animal imposed on an unsaturated soil is greater than the load‐bearing capacity of the soil. Compressive deformation or soil compaction is illustrated in Figs. 1 and 2. During compaction, particles are forced closer together by the applied load reducing the total pore space and permanently expelling air or water from the soil pores (Patto et al., 1978). This has a number of implications for soil hydrology and vegetation growth, as discussed in Section V.

246

G. S. BILOTTA ET AL. Before hoof force application Leaves of vegetation Roots of vegetation Soil aggregates collectively form the structure of the soil and support soil pores in between the larger aggregates

During hoof force application

Disturbed leaves of vegetation Pore space reduced The soil can become compacted to a depth of ~5 cm

After hoof force application Reduced vegetation cover and vigor

Shallow depression

Reduced macroporocity and infiltration capacity at the soil surface Increased soil bulk density and reduced porosity

Figure 1 A schematic diagram illustrating the process of soil compaction through the hoof force of grazing animals.

B. SOIL PUGGING Soil pugging is the term used to describe the process by which livestock tread on wet soft soil and create deep hoof imprints (Drewry, 2006). This is illustrated by Figs. 3 and 4. Pugging is a type of plastic deformation which occurs on soils with a medium soil moisture content when the animals’ load exceeds the bearing capacity of the soil (Patto et al., 1978). During plastic deformation, particles move relative to each other taking up new equilibrium positions with or without a reduction in total pore volume, although the proportion of fine pores will often be increased (Patto et al., 1978). The hoof imprints created by this deformation leave a very uneven pasture surface and can also influence soil hydrology and plant growth, as discussed in Section V.

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Figure 2 A photograph of a compacted soil under intensively managed grassland in Devon, United Kingdom. Photograph by G. S. Bilotta.

The factors controlling the amount of damage to pasture by pugging are the same as those for compaction (soil moisture content, stocking density, soil texture, vegetation cover) and the mechanisms by which these factors determine the amount of damage have already been discussed in the previous sections and so will not be reexamined. Pugging occurs at soil moisture contents intermediate to those at which compaction (low moisture) and poaching (high moisture) occur. However, soils with high clay contents can behave in a plastic manner even at lower soil moisture contents, making them particularly susceptible to pugging damage (Kellett, 1978).

C. SOIL POACHING Poaching is the term used to describe the slurry‐like soil conditions that occur on very wet soil when trampled by livestock (Drewry, 2006). This is illustrated in Figs. 5 and 6. Poaching is a type of elastic deformation which occurs when the animals load exceeds the load‐bearing capacity of the saturated soil and the hooves penetrate the soil surface (Kellett, 1978;

248

G. S. BILOTTA ET AL. Before hoof force application Leaves of vegetation Moist or wet topsoil Vegetation roots Soil aggregates

During hoof force application Plastic flow of soil surrounding the hoof Disturbed leaves of vegetation Moist or wet topsoil Soil aggregates break down and pore space is reduced. The soil may become compacted at some depth below the surface

After hoof force application

Reduced vegetation cover and vigor

Deep hoof print with raised edges remains in position following removal of the hoof

Reduced porosity and infiltration capacity at the soil surface Increased soil bulk density and reduced porosity

Figure 3

A schematic diagram illustrating the process of soil pugging by grazing animals.

Patto et al., 1978). Elastic deformation is associated with lateral bulging of the soil and usually occurs in soils with a high proportion of water‐filled pores (Patto et al., 1978). Since water is relatively incompressible, the soil recovers without an appreciable change in volume (Patto et al., 1978). In some cases, water held in the soil pores is forced from the soil as pressure is applied, but on removal of the load, the water is drawn back between the particles and elastic recovery occurs (Harris, 1971). Kellett (1978) suggests that the two major factors controlling the amount of damage to pastures as a result of poaching are: (1) the moisture content and (2) the stocking density. The mechanisms by which these factors determine the amount of damage have already been discussed in the previous

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Figure 4 A photograph of a pugged soil under intensively managed grassland in Devon, United Kingdom. Photograph by G. S. Bilotta.

section and so will not be reexamined. Poaching can be extremely disruptive to plant growth and also has serious implications for soil hydrology, as discussed in Section V.

V. THE IMPACTS OF SOIL STRUCTURAL ALTERATION BY GRAZING ANIMALS A. TREADING AND SOIL PHYSICAL PROPERTIES Changes to soil physical properties caused by grazing animals have received little attention compared with compaction of arable soils, despite the serious implications and the fact that in contrast to arable soils, there is little opportunity to ameliorate compacted soil through tillage in permanent pasture (Greenwood and McKenzie, 2001). A number of authors have reported that the eVect of stock treading and the resultant deformation tends to be confined to the upper layers of soil, within 50 mm of the soil

250

G. S. BILOTTA ET AL. Before hoof force application Vegetation leaves Saturated topsoil Vegetation roots Soil aggregates

During hoof force application

Elastic flow of soil surrounding the hoof

Hoof penetrates saturated soil surface

Disturbed leaves of vegetation Saturated topsoil Soil compacted at some depth below the surface

After hoof force application

Reduced vegetation cover and vigor

Relatively flat microtopography of saturated soil surface as a result of elastic flow following the removal of the hoof

Reduced macroporosity and infiltration capacity at the soil surface after formation of a surface pan layer Increased soil bulk density and reduced porosity

Figure 5

A schematic diagram illustrating the process of soil poaching by grazing animals.

surface (Alderfer and Robinson, 1947; Climo and Richardson, 1984; Drewry, 2006; Greenwood et al., 1997; Mulholland and Fullen, 1991; Tanner and Mamaril, 1959). Alderfer and Robinson (1947) argue that it is this zone which is very important in determining soil hydrology and plant growth/ vigor. One of the main eVects of soil compaction by grazing animals is the reduction in pore space and macroporosity which has often been associated with an increase in bulk density (Climo and Richardson, 1984; Di et al., 2001; Drewry and Paton, 2005; Mulholland and Fullen, 1991; Willatt and Pullar, 1983). The reduction in pore space is known to influence soil hydrology, vegetation growth, and the vitality of soil fauna, these impacts are discussed in the following sections. Soil pugging incorporates the eVects of soil compaction, as well as an alteration of soil surface microtopography, creating a

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Figure 6 A photograph of a poached soil under intensively managed grassland in Devon, United Kingdom. Photograph by G. S. Bilotta.

rough and uneven surface. Soil poaching is associated with a breakdown of soil aggregates and a rearrangement of soil particles in the surface soil horizons. This can lead to an increased bulk density and may also lead to the formation of a surface pan as the soil dries (Kellett, 1978). O’Connor (1956) suggests that it is also possible, when the soil moisture content is high, for hoof forces to greatly exceed the bearing strength of the soil and thus for hooves to penetrate the soil surface and cause large increases in soil bulk density at a greater depth in the soil, compressing the soil at some depth below the surface.

B. TREADING

AND

SOIL HYDROLOGY

Soil compaction and the reduction in pore space can lead to decreases in both the hydraulic conductivity of the soil (Greenwood et al., 1997; Willatt and Pullar, 1983) and the infiltration capacity of the soil (Alderfer and Robinson, 1947; Heathwaite et al., 1990; Mulholland and Fullen, 1991). This can make the soil more prone to ponding, thus rendering the soil

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more susceptible to further deformation such as poaching (Mulholland and Fullen, 1991; Patto et al., 1978). Furthermore, it can make the soil more prone to surface runoV generation as drainage becomes impeded. For example, Heathwaite et al. (1990) found that infiltration capacity was reduced by 80% and surface runoV volumes were increased by nearly 12 times on heavily grazed grassland compared with ungrazed grassland. Mulholland and Fullen (1991) showed that infiltration rate was very sensitive to the soil structural change caused by stock treading and showed a 98.5% decrease in infiltration in heavily trampled areas, although the largest decrease in infiltration occurred on initial compaction (87.5%), with only minor decreases for subsequent compactions. Soil compaction is not the only form of deformation that can influence soil hydrology. For example, soils that have undergone pugging may be more prone to ponding in surface depressions. Similarly, soils that have undergone poaching and the formation of a surface‐pan may also be more susceptible to ponding and the generation of surface runoV. These changes to the soil hydrology have implications for runoV from grazed land, potentially modifying not only the quantity of runoV, but also the quality of runoV, in terms of sediment and nutrient loads moving over and through the soil (Di et al., 2001; Heathwaite et al., 1990). The relevance of the changes in soil properties as a result of treading along with the eVects of other activities carried out by grazing animals is illustrated in Fig. 7. These changes to the soil physical properties can also influence nutrient transformation processes within the soil by altering the moisture regime and aVecting soil redox potential and plant uptake processes (Climo and Richardson, 1984; Di et al., 2001).

C. TREADING

AND

VEGETATION GROWTH

Several workers have reported that treading by grazing animals can cause a significant reduction in herbage growth/yield (Cluzeau et al., 1992; Edmond, 1962; Federer et al., 1961; Matches, 1992; Tanner and Mamaril, 1959). It has been estimated that the reduction in herbage yield may be as large as 25–40% (Carter, 1962; Muller, 1965; Schothorst, 1963), made up of 5–20% from immediately damaged and buried herbage, and 10–20% from reduced production by the remaining damaged sward (Kellett, 1978). Tanner and Mamaril (1959) reported a 20% decrease in herbage yields as a result of soil compaction which caused a severe decline in soil pore space and aeration in the root zone. Cluzeau et al. (1992) found that direct plant damage during stock trampling was responsible for the destruction of a large amount of plant material and a reduction in herbage yield. Although pasture grasses are particularly well adapted to frequent harvesting of their leaves, possessing condensed growing points that are close to the base of the plant, the action of

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Livestock (type, age, stocking density)

Treading

Defoliation

Soil Dry soil Moist soil

Decreased porosity Decreased infiltration

Increased infiltration-excess overland flow

Compaction

Increased bulk density

Decreased hydraulic conductivity

Pugging

Decreased macroporosity

Decreased soil shear strength

Increased saturationexcess overland flow

Increased overland flow

Direct effects (e.g., crushing, bruising, burial in mud)

Poaching Decreased herbage yield

Indirect effects

Minimal deformation

Vegetation

Wet soil Saturated soil

Decreased percentage cover Decreased biodiversity

Reduced protection of soil surface

Increased erosion

Water quality deterioration (e.g., sedimentation problems, eutrophication, bacterial contamination)

Figure 7 A conceptual diagram representing the environmental degradation induced by overgrazing (treading and defoliation).

poaching can damage these growing points and bury plant structures in mud (Kellett, 1978). In addition, grass roots can be seriously aVected by poaching as these tend to be concentrated in the top 50 mm of the soil—the zone of poaching damage (Kellett, 1978). However, it has been diYcult to determine the extent to which pasture response occurs due to changes in soil physical condition caused by soil compaction, pugging, and poaching alone, or due to trampling on the plant matter directly, leading to plant bruising, crushing, root damage, and plant displacement or burial in mud. A study by Drewry et al. (2001), which simulated dairy cow treading in controlled field conditions allowing for soil compaction, but not pugging and with minimal plant damage, found that significant reductions in pasture yield could be induced by compaction alone (i.e., indirect eVects). Nevertheless, in the real field situation, reductions in plant growth are probably a result of a combination of direct and indirect eVects of treading and soil compaction (Di et al., 2001). As well as the reduction in grassland vegetation yield due to treading eVects, the uptake of plant matter by grazing animals may also be reduced. This is because grazing animals will often refuse herbage when it is rendered unpalatable by contamination with mud or when herbage is broken and left lying on the soil surface (Kellett, 1978).

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D. TREADING AND SOIL FAUNA It has been noted that soil fauna generally have positive eVects on the soil by: (1) increasing the porosity and permeability, (2) improving soil structure, and (3) enhancing nutrient cycling and soil fertility (Trimble and Mendel, 1995). Earthworms (Lumbricina) are often viewed as the most beneficial fauna in terms of improvement of the soil structure (Piearce, 1984; Trimble and Mendel, 1995). However, some workers have indicated that soil deformation by grazing animals can cause a decline in earthworm numbers (Cluzeau et al., 1992; Drewry and Paton, 2005; Knight, 1992; Piearce, 1984). Earthworms have diYculty surviving in impacted soil conditions resulting from heavy grazing, although it has not yet been established whether this is due to direct impacts (i.e., mortality due to crushing) or indirect impacts (i.e., sublethal eVects due to environmental changes within the soil) (Cluzeau et al., 1992; Trimble and Mendel, 1995). Piearce (1984) proposed that there are several mechanisms for the reduction in earthworm numbers in trampled areas. First, trampling can cause death by direct crushing of individuals in the soil surface. Second, a less immediate but possibly equally important mechanism is associated with the reduction in soil porosity which impairs the movement of water and air through the soil and impedes earthworm locomotion. Third, trampling and defoliation can alter the amount of vegetation and hence the quality and quantity of food available for the soil fauna. Finally, the reduction in vegetation height and cover associated with trampling and defoliation can result in a harsher microclimate at the soil surface and diminished protection from predators. Furthermore, Piearce (1984) found no evidence to support the contention that higher numbers of dung pats in heavily trampled areas could counteract the eVect of reduced vegetation and soil deformation. Regardless of the causal mechanism, reductions in earthworm numbers as a result of grazing could lead to a loss of the important beneficial activities carried out by the organisms, which includes recovery of the soil after compaction (Drewry, 2006).

VI. THE IMPACT OF DEFOLIATION BY GRAZING ANIMALS ON GRASSLAND VEGETATION The vegetation of grasslands is central to the livestock/dairy production system. It provides forage for grazing animals and is often used as a source of food (in the form of hay or silage), while the animals are housed indoors over the winter period. Vegetation also protects the soil surface from the treading eVects of grazing animals as well as the erosional influence of rain‐ splash and surface runoV. This, in turn, can help preserve the water quality in surface waters within these environments. However, the high stocking densities

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associated with intensively managed grasslands can lead to excessive defoliation of vegetation and can have serious implications on pasture herbage yield, vegetation percentage cover, and vegetation biodiversity. Livestock can defoliate large amounts of vegetation while grazing. The response of the grassland vegetation to this defoliation will depend on factors such as: (1) the frequency and severity of vegetation removal/grazing, (2) the degree of compaction, pugging, and poaching on the soil, and (3) the amount of excreta deposited onto the pasture. These factors are determined primarily by the species and age of grazing animal, the stocking density, soil texture, soil moisture content, and farm management factors, which are discussed below.

A. ANIMAL SPECIES

AND

AGE

DiVerent species of animals graze diVerently in terms of how much they eat (quantity) and what they eat (selectivity). The quantity of vegetation consumed is a function of body mass and stage in the animal’s life cycle. The selectivity is a consequence of diVerences in the animal’s mouth size, lip anatomy, and method of prehension (Matches, 1992). Livestock are capable of consuming large amounts of vegetation while grazing. For example, estimates of the daily consumption by cows, based on UK studies, range from 7 kg dry matter per day for heifers (Rook et al., 2004; Rutter et al., 2002) to between 14 and 18 kg dry matter per day for dairy cows (Gibb et al., 1999; Orr et al., 2001; Rutter et al., 2004). Given that dry matter constitutes around 20% of the fresh weight of vegetation, one dairy cow is capable of ingesting 100 kg of fresh plant matter per day (Rook et al., 2004). This is clearly a significant quantity of herbage consumption, particularly when the mean net rate of herbage growth is considered to be 60 kg dry matter per hectare, per day for the United Kingdom intensively managed grasslands (Orr et al., 1988). Furthermore, this vegetation removal is not likely to be evenly distributed across a designated grazing area. Grazing livestock display both positive selection of desirable areas (fresh young grass/clover shoots) and avoidance of undesirable areas (e.g., dung pats, coarse grasses) (Rook et al., 2004), as well as tendencies to aggregate and spend a disproportionately large amount of time grazing and walking adjacent to fence lines. This can lead to the development of overgrazed patches, even to the extent of death of vegetation (Matches, 1992). Selectivity of grazing by some animals can even be used to isolate preferred species of plant and individual parts of plants. For example, Hughes et al. (1984) found that the diet of lambs (consuming 1.5 kg dry matter per day) contained a greater portion of clover (þ23%) and a smaller portion of ryegrass (19%) and of dead material (3%) than that of calves (consuming 7 kg dry matter per day), demonstrating some of the diVerences in grazing selectivity between species, and that lambs could be more selective while

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grazing than calves. The selectivity of grazing animals can be an important factor in determining the pasture species composition and the stability of the ecosystem.

B.

STOCKING DENSITY

As stocking density is increased, the frequency and closeness of defoliation increases (Matches, 1992). This leads to lower herbage availability per animal which in turn can cause the animals to be less selective in what they eat (Matches, 1992). Therefore, as stocking rate increases, the level of plant defoliation increases along with changes in sward morphology and composition (Matches, 1992). The ultimate economic goal of intensive livestock farming is to maximize production at minimum expense. Pastoral production can be increased either by raising production per animal (by using, for example, concentrated feeds, heated housing, and veterinary medicines), or by increasing the stocking density (Langlands and Bennett, 1973). However, there are natural limits on production per animal imposed by the rates of natural internal biochemical processes. There are also natural limits to stocking density due to the rate of herbage growth/production. While herbage yield can be enhanced via the use of fertilizers and weed‐control treatments, this enhancement cannot continue indefinitely. For decades, scientists have indeed considered stocking density in relation to the carrying capacity of the pasture vegetation (Evans, 1998). The carrying capacity of a pasture, in this case, is an estimate of how much and what type of vegetation will grow there, and in turn, how many grazing animals this vegetation can support (Evans, 1998; Matches, 1992). The optimum stocking density in terms of production is that at which the leaf area of vegetation is maintained at a level which allows for maximum growth rates throughout the grazing season. However, while the carrying‐capacity concept considers the consumption rate of vegetation by grazing animals and the threat of overgrazing to vegetation yield, it does not consider the other environmental impacts of grazing animals or how these are aVected by stocking density (Evans, 1998; Matches, 1992). Evans (1998) argues that the stocking density of a pasture will determine not only the total amount of vegetation consumed but also the total number of hooves impacting on the grassland soil surface, and the amount of excreta deposited onto the pasture, therefore, estimates of sustainable production should include these factors into stocking density planning. Overgrazing is the cause of 23% of the soil degradation in Europe (RCEP, 1996). Overgrazing is not a new phenomenon and there have been many cases of severe overgrazing in the United Kingdom in the past 200 years (Johns, 1998).

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C. VEGETATION RESPONSE One of the most important impacts of defoliation is associated with the change in herbage yield. Changes to the productivity of grassland vegetation in response to grazing depend strongly on grazing intensity/stocking density. McNaughton (1983, 1986) and Hodgkinson and Mott (1986) have proposed three alternative hypotheses on how plant growth and fitness may respond to grazing intensity. Response one is where net primary productivity (NPP) of plants shows a consistent decline as the intensity of grazing increases. This is probably the most common view among ecologists and evolutionary biologists; it is based on the principle that herbivory is always detrimental to the plant eaten. Edmond (1958, 1962) reported this type of response from studies on perennial ryegrass and white clover pasture in New Zealand. The herbage yield from pasture (maintained in a wet state) with stocking densities of 3, 6, 9, 12, and 18 sheep per hectare were 808, 644, 490, 267, and 127 kg dry mass per 0.405 hectare, respectively (Edmond, 1962). Response two is where the plants are able to compensate for tissue removal up to some level, beyond which plant productivity begins to decline as the intensity of defoliation increases further. Langlands and Bennett (1973) found this type of response from their five‐year study on the eVect of sheep stocking density on pastoral production in New South Wales, Australia. In this example, herbage production was relatively insensitive to increases in stocking density over the range of 2–22 sheep per hectare, but beyond this range herbage productivity began to decline at a greater rate with increasing stocking density. Langlands and Bennett (1973) attributed the decline in productivity at the higher stocking rates to the decline in basal cover and expansion of bare areas of soil. Response three, perhaps the most interesting and controversial, is where moderate levels of defoliation may result in overcompensation by the plant, due to intrinsic or extrinsic consequences of defoliation (McNaughton, 1983). Thus, within some levels of defoliation, plant productivity may be enhanced. McNaughton (1983) termed this as ‘‘overcompensatory growth.’’ Indeed, the principle that certain levels of defoliation may enhance and stimulate herbage growth has long been used in the production of turfgrasses (whereby frequent clipping at a moderate height is used) (Albert, 1927; Mortimer and Ahlgren, 1936). An example of this type of response was reported by Vickery (1972) who found that NPP was greatest at 20 sheep per hectare and least at 10 sheep per hectare, while NPP at 30 sheep per hectare was greater than at 10 sheep per hectare. Vickery (1972) attributed the lower NPP at 10 sheep per hectare to reduced photosynthesis as a result of canopy closure and increased plant competition in the absence of regular grazing. The highest NPP at 20 sheep per hectare was attributed to increased photosynthetic eYciency because of a higher proportion of younger tillers and plants at this stocking density (Vickery, 1972). The decline in NPP at the highest stocking density was attributed to the reduced

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plant material available for photosynthesis as a result of excessive herbage consumption. McNaughton (1979) argued that simplified statements about the eVect of damage to vegetative tissue on the ultimate yield of those or other tissues, implying a uniform monotonic series of eVects on the plant by an increasing series of herbivory levels, are highly misleading. Rather the plant responds to a whole complex of environmental factors, and the variety of plant responses are subject to the constraints of plant genetics, developmental stage, and the plant tissues that are aVected (McNaughton, 1983). Nevertheless, one key point which holds true in all of the above hypothetical responses is that at high levels of herbivory and defoliation plant productivity is negatively aVected. Therefore, overstocking of grasslands may reduce livestock yield as well as having an impact on the plant system. If intensively managed grasslands are to be maintained in a productive yet environmentally friendly state, then a complete understanding of the eVect of grazing on vegetation is critical. At present, however, a review of the literature reveals gaps in our knowledge regarding the responses of grassland vegetation to defoliation by grazing animals.

VII. IMPACT OF EXCRETION BY GRAZING ANIMALS ON VEGETATION, SOILS, AND SURFACE WATERS IN INTENSIVELY MANAGED GRASSLANDS Livestock can produce large quantities of waste (urine and feces), with dairy cattle being the highest producers. For example, in the European Union, there are in excess of 24 million dairy cattle (Eurostat, 2006), each adult cow producing an average of ca. 20 tons of slurry (a mix of urine and feces) each year (Smith and Frost, 2000). On an annual basis, 50% of this excreta is voided in the field while grazing (Chadwick and Chen, 2002), with the majority of the remainder being collected in the form of manure or slurry when the animals are housed indoors over the winter period (Mawdsley et al., 1995). The waste collected while the animals are housed indoors may also eventually be applied to the pasture surface through slurry spreading or manure application. This waste is a source of organic matter and nutrients [nitrogen (N) and phosphorus (P)] and is also a potential source of pathogens.

A. LIVESTOCK WASTES

AS A

SOURCE

OF

NUTRIENTS

Livestock wastes are often a rich source of nutrients such as N and P because only a small percentage (3–30%) of the nutrients in the food ingested by the animal is actually utilized by the animal and assimilated into its tissues,

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the remainder being excreted in feces and urine (Holmes, 1970; Tamminga, 1992). The nutrient content of livestock excreta may be enhanced further when the animals have been fed concentrated feeds (Tamminga, 1992). While the nutrient and organic matter content of animal waste can be considered to be beneficial to plant growth, long‐term fertility, and soil structure in grassland environments, accumulation of nutrients in grassland soils has been shown to cause a shift in grassland plant diversity and botanical composition which has subsequent eVects on the insect and invertebrate communities (Matches, 1992). Generally, the addition of N in urine stimulates growth of dominant grass species (such as L. perenne) and the addition of P in dung stimulates the growth of dominant legume species (such as T. repens) especially on P‐deficient soil (Matches, 1992). In addition, plants immediately beneath dung pats may be killed due to absence of light, and urine occasionally scorches the sward (Matches, 1992). Furthermore, the accumulation of these nutrients in the soil and resultant increased delivery of excessive amounts of these nutrients (particularly N, P, and C) into surface waters is associated with eutrophication problems such as the growth of toxic algal blooms which pose a threat to the health of humans and domesticated and wild animals (Chadwick and Chen, 2002; Haygarth and Jarvis, 1999). Defecation by grazing animals can also influence the distribution of nutrients in the soil and the spatial pattern of nutrients across the pasture. First, dung deposition on grassland tends to lead to higher concentrations of nutrients in the surface soil horizon in the absence of ploughing (Haygarth et al., 1998). Second, there may be spatial excreta deposition hotspots associated with the movement behavior of livestock and the tendency to concentrate in camping‐grounds or in sheltered spots overnight. These hotspots may be significant sources of readily available sediments and colloids, N and P, and pathogens, posing a high risk to surface water quality where these hotspots coincide with surface runoV flow paths (Page et al., 2005), which may be promoted by the compaction and soil deformation eVects of stock treading.

B. LIVESTOCK WASTES

AS A

SOURCE

OF

PATHOGENS

The rumen and digestive tract of agricultural livestock is host to a rich diversity of microflora and can also act as a reservoir for pathogenic (disease‐ causing) microorganisms (Rasmussen et al., 1993). Some of these pathogens are excreted in the feces of infected, and in some cases, healthy ‘‘carrier’’ animals (Chadwick and Chen, 2002). While some pathogens are obligate parasites and are of limited concern, others can survive saprophytically in the environment for long periods and pose a threat to other organisms (Mawdsley et al., 1995). It has been suggested that modern intensive grassland management practices are contributing to greater abundance and survival of

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pathogens in livestock wastes. For example, prior to agricultural intensification, housed livestock were often bedded on large amounts of straw and the waste was managed as farmyard manure (Jones, 1982). Traditionally, this farmyard manure was composted, an aerobic process where temperatures often rise as high as 70 C and therefore the majority of pathogens were destroyed (Jones, 1980). However, as herd size and the number of housed animals has increased, there has been a move toward the collection of waste in a semiliquid, slurry form which contains only a minimum amount of solid bedding material (Mawdsley et al., 1995). It is estimated that 50–60% of waste from housed cattle is now managed as slurry (Smith and Unwin, 1983). However, in intensive systems, slurry is collected and stored under conditions which rapidly become anaerobic and hence temperature rise and the concurrent destruction of pathogens, seen in composting, does not occur (Rankin and Taylor, 1969). This slurry, containing pathogens, is often applied to the pasture surface where it potentially may be washed oV into surface waters during rainfall events. The pathogens in livestock waste which pose the greatest threat to human health are bacterial pathogens such as Eschericheria coli O157 and Salmonella spp., viruses such as Rotavirus spp., and protozoa such as Cryptosporidium and Gardia spp. (Mawdsley et al., 1995). These pathogens can be transferred to surface and drinking waters via hydrological transport in association with colloidal matter present in dung and soil (Chadwick and Chen, 2002).

VIII. IMPACTS OF GRAZING ANIMALS ON THE WATER QUALITY OF SURFACE WATERS IN INTENSIVELY MANAGED GRASSLANDS The impacts of grazing animals on surface waters, unless livestock are allowed direct access to the channel network, are very much secondary eVects resulting from the impacts of the grazing animals on the soils and vegetation of the grassland. This chapter has been structured around the principle that livestock carry out three key activities which may cause environmental degradation in intensively managed grassland environments: (1) treading, (2) defoliation, and (3) excretion. The mechanisms by which these activities can cause degradation of grassland soils and vegetation have already been discussed in previous sections; however, the ways in which this degradation can be transferred to surface waters have not yet been fully discussed. This section of the chapter examines the potential means by which grazing animals can indirectly impact on surface waters in grassland environments, as is illustrated in Fig. 8. This section is divided into the following parts: (A) soil erosion and sedimentation problems, (B) eutrophication, and (C) pathogenic

Livestock (type, age, stocking density)

Increased infiltration-excess overland flow

Decreased macroporosity

Decreased hydraulic conductivity Increased saturationexcess overland flow

Increased overland flow

Decreased herbage yield

Decreased soil shear strength

Decreased percentage cover

Reduced protection of soil surface

Manure

Application

Poaching

Slurry

Organic matter

Pathogens (e.g., E. coli)

Nutrients (e.g., N, P, K)

Storage Decreased biodiversity

Leakage

Decreased infiltration

Increased bulk density

Pugging

(e.g., crushing, bruising, burial in mud)

Indirect effects

Decreased porosity

Compaction

In pasture

Direct effects

Dry soil Moist soil Wet soil Saturated soil Minimal deformation

Excretion In housing

Vegetation

Soil

Direct runoff/incidental transfer

Increased erosion

Immobilisation by soil microbes

Sorption to soil colloids Water quality deterioration (e.g., sedimentation problems, eutrophication, bacterial contamination)

Uptake by plants

Incorporation into soil structure by soil fauna

Erosion

Figure 8 A conceptual model illustrating the impacts of treading, defoliation, and excretion by grazing animals on the water quality of surface waters in intensively managed grasslands.

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Defoliation

Treading

261

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contamination. These divisions reflect diVerences in the level of knowledge and understanding relating to each area.

A. SOIL EROSION

AND

SEDIMENTATION PROBLEMS

Soil erosion can be defined as the removal of soil by wind, water, and mass movements at a faster rate than at which new soil forms (Morgan, 1980). While the rates of soil erosion in temperate regions are likely to be relatively small in comparison with those in tropical‐humid regions, the impacts of soil erosion in temperate regions are certainly not insignificant when considering the environmental and economic costs which are incurred as a consequence of this process (Morgan, 1980). For example, the estimated cost of soil erosion to the UK economy is around £90 million per annum (Environment Agency, 2002). Furthermore, the costs of cleaning up the polluting impacts of soil erosion on water in the United Kingdom are estimated at £260 million per annum (Evans, 1995). The costs of soil erosion are therefore a result of both on‐site and oV‐site impacts. On‐site impacts are particularly important on agricultural land where they can lead to the redistribution of soil within a field, the loss of soil from a field, the breakdown of soil structure, and the decline in organic matter and nutrients. This, in turn, results in a reduction of cultivable soil depth and a decline in soil fertility (Morgan, 2005). OV‐site problems arise from the delivery of eroded sediment into surface waters, which reduces the capacity of rivers and drainage ditches, enhances the risk of flooding, blocks irrigation canals, and shortens the design life of reservoirs (Morgan, 2005; Verstraeten and Poesen, 2000). The cost of damages and dredging stream channels as a result of soil erosion from agriculture in the United Kingdom is estimated at £7.8 million per annum (Environment Agency, 2002). Sediment delivery to surface waters can also have direct ecological impacts by, for example, interfering with fish spawning/incubation sites (Greig et al., 2005; Walling et al., 2003). Salmon and trout in north‐west Europe have declined from great abundance in preindustrial times to the present day, where they are absent from former habitats or where the threat of extinction to sparse residual populations is real (Harrod and Theurer, 2002). Causes of their reduction may be numerous and complex, but sediment intrusion into spawning gravels is one of them and has potential to cause serious damage to fish stocks (Greig et al., 2005; Harrod and Theurer, 2002; Walling et al., 2003). Sediment delivery to surface waters can also have indirect ecological impacts because these particles can act as vectors of sorbed contaminants such as pesticides (Morgan, 2005), pathogens (Oliver et al., 2005a) and P (Fraser et al., 1999; Sharpley

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et al., 1994; Svendsen et al., 1995). The sediment‐facilitated transport of P to surface waters is a particularly significant threat to water quality because of the role that P plays in eutrophication. While authors in the ‘‘grazing impacts’’ literature have often acknowledged and recognized that the changes to soil physical properties, vegetation cover, and pasture hydrology brought about by intensive grazing may have implications for increasing soil erosion, very little quantitative data exists to directly support this hypothesis from intensively managed grasslands. In fact, grasslands as a whole have largely been ignored as potential sources of sediments by the soil erosion community (Evans, 1998; Heathwaite et al., 1990; Heathwaite and Dils, 2000). In the United Kingdom particularly, the quantification of soil erosion has focused on the arable sector of the agricultural land (Boardman, 2002; Brazier et al., 2007; Evans, 1997; Morgan, 2005). This research bias has, in the past, been justified by the belief that grasslands do not yield significant amounts of sediment due to the eVect of their high surface cover which acts to intercept raindrops and retard runoV, resulting in limited detachment and transport of soil particles (Nash and Halliwell, 1999; Nash and Murdoch, 1997). This chapter has highlighted several reasons why this perception may be false, particularly in intensively managed grasslands where (1) soil compaction, pugging, and poaching can promote surface runoV generation and reduce the resistance of the soil to erosion, (2) defoliation and direct and indirect treading eVects can dramatically reduce the protective vegetation cover, and (3) feces deposited onto the pasture by grazing animals and slurry/manure applied to the pasture by farmers provide a readily available source of particulate material. However, no new research has proven these links as yet, so they remain largely anecdotal and theoretical. Evans (1997) recommends that further research is needed into the erosion initiated by grazing animals and recommends that national surveys of erosion by grazing animals should be conducted. This information is vital if governments are aiming to be able to eVectively mitigate water quality issues in surface waters. Failure to recognize the potential sources of surface water pollutants in catchments, and continuation of a ‘‘perceptual understanding’’ of the factors that control the magnitude of these sources (rather than an empirical understanding), will inevitably result in failure of water‐quality remediation measures.

B. EUTROPHICATION Eutrophication is a process of nutrient enrichment which increases the primary productivity of surface waters and can potentially impact on all water types ranging from those that are nutrient poor (oligotrophic) to those

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considered to be nutrient enriched (eutrophic) (Foy, 2005). Eutrophication is a slow and benign natural process associated with the ageing of a lake or waterbody; however, it can be accelerated and become harmful to ecosystem health if, for example, the anthropogenic input of nutrients occurs. The nutrients that are commonly of particular concern with regards to freshwater eutrophication are N and P. Phosphorus is most often the limiting nutrient in freshwater aquatic systems and is therefore commonly the prime cause of eutrophication (Sharpley and Rekolainen, 1997; Sharpley and Smith, 1990). The significance of P is evident from the strong correlation between mean total P (TP) measured in lakes over a wide geographic area and chlorophyll a, which is used as a surrogate for algal abundance (Canfield, 1983; Forsberg and Ryding, 1980; McCauley et al., 1989; OECD, 1982; Prairie et al., 1989; Pridmore et al., 1985; Seip et al., 2000; Smith, 1998). Even relatively low concentrations of P in surface waters can lead to eutrophication problems. The Organization for Economic Cooperation and Development (OECD) suggests that eutrophication problems can be triggered by P concentrations as low as 35–100 mg liter1 (OECD, 1982). For oligotrophic–mesotrophic waters, excessive P inputs can result in an increase in fish size which may be considered to be beneficial to some lake users, but in biodiversity terms can be damaging if other fish species are in decline (Foy, 2005). For eutrophic waters, the input of P may maintain or exacerbate a range of undesirable eVects. Eutrophication of waters can cause problems with its use for fisheries, recreation, industry, and drinking. For example, P can promote the excessive growth of aquatic vegetation and algae. The senescence and decomposition of this matter can deplete water oxygen levels which may lead to the mortality of fish and other aquatic organisms (Heathwaite, 1994). In addition, the cyanobacteria or blue‐green algae commonly associated with eutrophic waters present particular water‐ quality problems, as some species produce fast acting neurotoxins and slower acting hepatotoxins which can have serious adverse impacts on the health of humans and domesticated and wild animals (Foy, 2005). For example, at Rutland Water in Leicestershire (United Kingdom) in 1988, the bacteria Clostridium botulinum, which flourishes in anoxic sediments, developed to the extent that it caused botulism in birds and mammals (including domestic pets) using the lake (Heathwaite, 1994). In addition to toxins, cyanobacteria and some algae can produce other dissolved organic compounds (DOCs), principally geosmin and isoburneol, that can cause taste and odor problems (Cooke and Kennedy, 2001; Watson et al., 1999). If chlorine reacts with the DOCs formed by algal cell lysis or algal extraction during water treatment, potentially carcinogenic trihalomethanes can enter potable water supplies (Hoehn et al., 1980). Phosphorus, as the most common limiting nutrient in freshwater ecosystems, plays a key role in determining the presence of such harmful substances

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(e.g., neuro‐ and hepatotoxins and DOCs) in surface waters. There are several ways in which intensively managed grasslands can contribute nutrients to surface waters, and thus contribute to eutrophication. (1) RunoV of dissolved and particulate nutrients found in animal excreta which was deposited on the pasture while the animal was grazing. (2) RunoV of dissolved and particulate nutrients found in animal excreta which was applied to the pasture surface in the form of slurry and/or manure (collected while the animals were housed indoors). (3) RunoV of dissolved nutrients found in fertilizers which were applied to the pasture to enhance the pasture herbage yield. (4) Erosion and delivery of nutrients that are sorbed to soil particles and colloids. Mechanisms (1), (2), and (3) have received a fair amount of research attention over the last decade (Chardon et al., 1997; Dougherty et al., 2004; Edwards and Withers, 1998; Foy, 2005; Hart et al., 2004; Haygarth and Jarvis, 1999; Haygarth et al., 1998; Heinonen‐Tanski and Uusi‐Kamppa, 2001; Hooda et al., 2001; Nash and Halliwell, 1999; Preedy et al., 2001). However, as discussed in Section VIII.A, the potential for mechanism (4) to operate in intensively managed grasslands has largely been overlooked.

C. PATHOGENIC CONTAMINATION Impairment of waterways and receiving lakes by pathogenic pollution has a significant impact on human health and quality of life, with contamination of drinking water supplies and the closure of recreational surface waters being two common consequences (Jamieson et al., 2005). The pathogens which pose the greatest threat to human health are bacterial pathogens such as E. coli O157 and Salmonella spp., viruses such as Rotavirus spp., and protozoa such as Cryptosporidium and Gardia spp. All of these pathogens can often be found in livestock wastes (Mawdsley et al., 1995), consequently, livestock‐based agriculture is one of the main nonhuman sources of this kind of water pollution (Vinten et al., 2004). The pathogens contained in livestock wastes may enter surface waters (1) directly by leakage of wastes held in buildings or stores to drainage systems, (2) indirectly following the application of waste to land, or (3) indirectly from feces deposited onto the pasture while the livestock are grazing (Aitken, 2003; Oliver et al., 2005b; Rodgers et al., 2003). However, despite the serious implications of this type of surface water pollution, many aspects of bacterial survival and transport are poorly understood (Jamieson et al., 2005). A review by Mawdsley et al. (1995) highlighted, in particular, the lack of direct information on the movement of pathogenic microorganisms present in livestock waste through the landscape to surface waters, although it is known that bacteria are often transported in association with fine sediment and colloids (suggesting that soil erosion may be an important mechanism involved in the transfer process).

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IX. ENVIRONMENTAL DEGRADATION BY GRAZING ANIMALS: RECOVERY AND REMEDIATION A.

NATURAL RECOVERY OF SOIL PHYSICAL CONDITION FOLLOWING TREADING DAMAGE

To a certain extent, some of the damage to soil physical condition induced by grazing animals can be reversed by natural processes. Natural recovery of soil physical properties has been shown to be cyclical in nature, associated with wetting and drying cycles, subsequent soil cracking, earthworm burrowing, root penetration and decay, and freeze–thaw cycles during the winter (Dexter, 1991; Drewry, 2006; Drewry and Paton, 2005; Greenland, 1981; Greenwood and McKenzie, 2001; Hodgson and Chan, 1984). The time taken for natural recovery of soil physical condition varies depending on soil type, extent of initial damage, management methods, and climate, but may take anything from weeks to months, or even years (Drewry, 2006). A study by Drewry et al. (2004) found that most of the soil damage that occurred on a dairy farm in New Zealand took place in the wet spring and recovery of the soils physical condition occurred over the summer and autumn months, while recovery in the winter was much lower. The potential for natural recovery of soil physical condition needs to be balanced with the potential for further soil physical deterioration when regrazed and so relies heavily on land management factors (Drewry, 2006).

B.

MITIGATION

AND

DAMAGE REDUCTION METHODS

Damage to the soil by grazing animals can never be entirely avoided, but the damage can be minimized by intelligent land and grazing management (Kellett, 1978). This section of the chapter focuses on mitigation and damage reduction measures for reducing the impact of grazing animals. In this chapter, the measures have been divided into three categories: (1) livestock management, (2) land management, and (3) waste management. Examples of some of the methods from each category are discussed below and can be found in Table I.

1.

Livestock Management

Perhaps one of the most obvious methods for reducing the amount of damage to grassland vegetation and soil physical condition is that of reducing the total stocking density. Many authors suggest this as a remediation strategy (Kellett, 1978; Langlands and Bennett, 1973; Mulholland and Fullen, 1991;

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Table I Remediation and Mitigation Measures for Minimizing the Impacts of Grazing Animals Method Reduce stocking density Move livestock into housing or hard standings when the soil is wet or saturated Move grazing animals to drier areas of the pasture during wet periods Move grazing animals to sacrifice enclosures within the pasture when the soil is wet or saturated Increase hoof contact area through the use of livestock shoes Reduce the length of the grazing season Reduce dietary N and P intake from animal feeds Relocate feeding and drinking troughs in the pasture at regular intervals Install subsurface drainage Loosen compacted soil layers in grassland fields Terrace slopes Tillage and reseed Increase storage capacity of slurry and manure stores Switch from slurry to manure handling Adopt a batch storage method for slurry and manure storage Avoid applying slurry onto high risk areas and at high risk times Inject slurry into the soil rather than spreading it with a splash plate Integrate fertilizer and manure/slurry nutrient supply

References Patto et al. (1978) Kellett (1978) Sears (1956) Mulholland and Fullen (1991) Wind and Schothorst (1964) Davies and Armstrong (1986) Tamminga (1992) Hilton (2002) Armstrong and Garwood (1991); Davies and Armstrong (1986) Harrison et al. (1994) Gassman et al. (2006) Johnson et al. (1993) McGechan and Wu (1998) Mawdsley et al. (1995) Chadwick and Chen (2002) Haygarth and Jarvis (1999) Hilton (2002) Unwin et al. (1986)

Patto et al., 1978; Willatt and Pullar, 1983). This measure works in a number of ways. First, it reduces the number of hooves and the frequency that hooves impact on the pasture surface which in turn reduces the amount of soil deformation (Patto et al., 1978), and damage to vegetation (Di et al., 2001; Kellett, 1978). Second, it reduces the frequency and closeness of defoliation (Matches, 1992), which in turn reduces the occurrence of bare patches and promotes the development of a healthy vegetation cover which acts to protect the soil surface (Evans, 1997). Third, it decreases both the amount of excreta deposited onto the pasture while grazing and the amount of excreta collected while the animals are housed indoors (which may be spread onto the field at a later date). This in turn lowers the potential for N, P, and pathogens (found in animal waste) to be transported from land to surface waters. However, while this measure would

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be relatively simple to implement, it would result in a reduction in farm income which could threaten the economic viability of production. A less drastic change for livestock management would be to move livestock into housing or onto hard standings when the grassland soil is wet or saturated (Kellett, 1978). This method is based on the principle that the resistance of a soil to deformation under treading declines as soil moisture increases and therefore the greatest amount of soil damage occurs when livestock tread on wet soils (Climo and Richardson, 1984; Patto et al., 1978; Wind and Schothorst, 1964). If livestock are removed from the pasture during these high risk times, damage to soils and vegetation will be limited. Farmers can maintain a regular assessment of the weather forecast and act accordingly. Preventing treading while rain is occurring and water is ponding on the soil surface has been suggested as a very simple but particularly eVective method for reducing poaching damage (Scholefield and Hall, 1985). However, this dynamic form of management is not always possible as housing and hard standings may already be in use or may not even exist on some farms. Perhaps a more practical alternative to this would be to move livestock to drier areas (if present) of the pasture during wet periods (Sears, 1956), or to move livestock into sacrifice enclosures of the pasture where they are allowed to damage only a small area (Mulholland and Fullen, 1991). With the latter method, the farmer would need to ensure that these sacrifice areas are poorly connected to the channel network to prevent eroded sediment and colloidal material from entering surface waters. Another measure which works by the same mechanism as above is that of reducing the length of the grazing season. In many temperate countries, precipitation is seasonal with the grazing season fitting between two hydrological seasons. However, with this livestock management method, there is a risk of animals being on the pasture while it is still wet (spring), or becoming wet (autumn). By reducing the length of the grazing season, the chance of livestock treading on wet soils is reduced (Davies and Armstrong, 1986). One issue with this damage reduction method is that it requires a greater amount of stored food (silage, hay, concentrated feeds) while the animals are housed indoors. These come at a cost to the farmer and can result in increased external N and P being brought into the grassland system. A novel suggestion by Wind and Schothorst (1964) involved the use of a type of shoe for livestock. This method is based on the principle that the forces imposed on the soil (which cause compaction, pugging, and poaching) by animal hooves are a function of animal mass and the surface area of the hooves in contact with the ground. By increasing the surface area of animal hooves (through the use of a type of shoe), the forces imposed on the soil are reduced and less deformation will take place as a result. Wind and Schothorst (1964) propose that relatively strong shoes could be made in bulk for little expense. However, while this suggestion makes good logical sense, there is little evidence to show that the shoes would actually be successful with real livestock.

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A method for reducing N and P contents of animal excreta and therefore N and P accumulation in soils and runoV to surface waters is that of reducing dietary N and P intake. In modern intensively managed grasslands, the diet of livestock is often supplemented with concentrated feeds which are enriched with N and P in excess of the animals’ requirements. This maintains high yields from the animals; however, any N and P in excess of animal requirements is excreted by the animal (Holmes, 1970; Tamminga, 1992). By reducing the N and P content of livestock feeds, this reduces N and P concentrations in excreta, reducing N and P accumulation in grassland soils and reducing the potential for N and P transfer to surface waters where they can contribute to eutrophication problems.

2. Land Management One large‐scale and long‐term land management strategy for the reduction of damage to soil physical condition by grazing animals is the installation of subsurface drainage (Armstrong and Garwood, 1991; Davies and Armstrong, 1986; Kellett, 1978; Mulholland and Fullen, 1991). The installation of subsurface drainage leads to a lowering of the water table level and a decrease in the moisture content of the surface soil, thus increasing the shear strength of the soil and its resistance to damage (Patto et al., 1978). Several workers have reported that soil damage is reduced when the water table is lowered from the surface, and is normally avoided where the water table is kept below 500 mm of the surface (Davies and Armstrong, 1986; Kellett, 1978; Patto et al., 1978). However, calculations of the pipe drain spacing required for an eVective drainage scheme on a typical grassland site in the United Kingdom, for example, reveal that pipes would need to be too close to be economically justified (Davies and Armstrong, 1986; Kellett, 1978). A less costly method of achieving this desired drainage involves the use of mole drainage over wide‐ spaced lateral pipes. This has been traditionally used in only a few grassland areas (Armstrong and Garwood, 1991), but where it has been used, it has been very successful at improving water removal from the surface soil and reducing damage to the soil by grazing animals (Davies and Armstrong, 1986; Kellett, 1978). However, while pipe and mole drainage has been proven to control soil damage on soils with a high clay content (Davies and Armstrong, 1986), it is not appropriate for soils with a clay content of <30% because of instability and collapse of the channels. Furthermore, there is uncertainty over how the installation of subsurface drainage influences sediment and nutrient transfers from land to surface waters and there is evidence to suggest that subsurface drains may, in some cases, act as preferential pathways for runoV and provide a direct conduit to watercourses (Chapman et al., 2001; Dils and Heathwaite, 1999; Heathwaite et al., 2005; Simard et al., 2000).

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A shorter‐term land management method for mitigating soil and vegetation damage, such as compaction, pugging, and poaching, involves tillage and reseeding (Johnson et al., 1993). This process breaks up layers of compacted soil and removes deep hoof prints, allowing vegetation to reestablish. However, a review of the literature on the eVects of tillage by Greenwood and McKenzie (2001) revealed variable responses. Dexter (1991) suggested that although a compacted grassland soil can be temporarily turned into a soil with apparently near‐perfect structure by tillage (e.g., a seed bed of 1‐ to 5‐mm diameter aggregates overlying a loosened, well‐drained subsoil), the structure produced in this way may be far from equilibrium, may be mechanically unstable, and may collapse when wet to be as bad, if not worse, as before tillage. Furthermore, tillage can accelerate the deterioration of the soil physical condition by, for example, accelerating decomposition of organic matter and by disrupting stable soil aggregates (Dexter, 1991). Tillage and reseeding is probably best used over small areas such as old sites of drinking troughs/ feeding troughs and gateways which have been moved, where the soil requires loosening and rejuvenating (Harrison et al., 1994). If tillage and reseeding is used over whole fields or significantly large areas, there is a risk of enhanced erosion if rainfall occurs shortly after tillage and before the vegetation has established. An alternative, more dynamic form of land management involves the regular movement of drinking and feeding troughs. The soil and vegetation around these features tend to receive the greatest damage by grazing animals because animals tend to congregate around them and so they are exposed to more frequent treading, defoliation, and defecation and so eventually become compacted, pugged, or poached, devoid of vegetation, and rich in excreta. These areas can then become critical source areas (CSAs) for sediment, N, P, and pathogens, threatening water quality in surface waters. Moving the troughs before significant visible degradation occurs can reduce pasture damage and minimize environmental consequences (Hilton, 2002). Similarly, gateways also receive a higher frequency of animal traYc than the rest of the pasture and so can become degraded areas and eventually CSAs for surface waters. Farmers need to make sure that these features are located away from the channel network. Alternatively, connectivity to the channel network can be reduced through the use of landscape features such as buVer strips and/or hedges (Hilton, 2002).

3.

Waste Management

As mentioned previously, livestock can produce large quantities of waste (urine and feces). Approximately 50% of this waste is collected while the animals are housed indoors and is stored in the form of slurry (a liquid

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mix of urine and feces) or manure (a heap of solid dung and bedding material). This waste can be recycled to land at some point and if used eYciently and eVectively can provide a host of benefits to both farmers and the environment (Oliver et al., 2005a). For example, recycling livestock waste to land prevents excessive fecal waste accumulation in farm storage systems (Unwin et al., 1986), builds soil quality, and returns valuable nutrients back to the grassland system (Hooda et al., 2000). However, if poorly managed, livestock wastes can present a threat to environmental quality if, for example, farm storage systems leak or overspill into streams or if runoV from grassland contains recently applied animal wastes which enter surface waters. One major factor contributing to poor waste management is related to the storage capacity of farm waste facilities. For example, farmers are often under pressure to apply slurry to in unsuitable conditions during the winter period, due to limited farm storage facilities (Chadwick and Chen, 2002; Edwards and Withers, 1998). There is much anecdotal evidence of slurry applications that occur in defiance of good agricultural practice, including slurry spreading onto wet or frozen land, and the ejection of slurry from roadways and farm tracks onto hillslopes of adjacent fields (Preedy et al., 2001). These practices are likely to promote pollution of surface waters. A simple remediation measure for this issue would be to increase the capacity of farm waste stores (McGechan and Wu, 1998). This would relieve pressure on farmers to empty their stores and apply waste in unsuitable conditions. Clearly, the enlargement of stores will come at a cost to the farmer. However, costs can be minimized if the farmer shifts to collecting waste in the form of solid manure rather than slurry, as it will not always be necessary to construct purpose built stores for manure handling. Solid manure can be stored temporarily in heaps on hard‐standings or at suitable locations within the farm and therefore the storage capacity for manure is inherently more flexible than that used for slurry (a fixed volume pit, tank, or reservoir). Another advantage of solid manure handling, as opposed to slurry collection, is associated with the reduced survivorship and abundance of pathogens (Jones, 1982; Mawdsley et al., 1995). As mentioned previously, the aerobic composting process can produce temperatures of up to 70 C which can kill oV many pathogens (Jones, 1982). In contrast, slurry storage leads to anaerobic conditions and therefore temperatures are not raised high enough to kill oV as many pathogens. An alternative method for reducing the number of pathogens in stored animal wastes is that of adopting a batch storage method for both slurry and manure storage, whereby only the older waste is applied to the land, while the fresher waste is being collected in a diVerent batch until it has decayed/ composted further. This decreases the number of pathogens applied to the land due to the pathogen die‐oV process during storage (Oliver et al., 2005a). A further improvement for reducing runoV of applied animal wastes from

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land to surface waters involves a method of injecting the slurry into the soil rather than spreading it over the surface (Heinonen‐Tanski and Uusi‐Kamppa, 2001). Heinonen‐Tanski and Uusi‐Kamppa (2001) found that injecting the slurry rather than spreading it on the surface, reduced total P and total N in runoV from grassland by an average of 81% and 73%, respectively.

X.

FUTURE RESEARCH

One of the defining characteristics of intensively managed grassland is the high stocking density. Authors often cite the stocking density of a pasture to be one of the most important factors controlling the magnitude and extent of environmental degradation. The stocking density determines the number of hooves and frequency of hooves impacting on the soil surface, amount, frequency and closeness of defoliation, and the amount of excreta deposited onto the pasture. At low stocking densities, grazing can be beneficial to the environment, enhancing nutrient cycling and promoting biodiversity. However, at high stocking densities, damage to the pasture may occur, threatening the sustainability of farming and potentially impacting on water quality in surface waters. One simple remediation measure for the reduction of this environmental degradation would therefore be to reduce livestock stocking densities to a more optimum level whereby environmental degradation is limited but economic viability is maintained. The question is, ‘‘what is this optimum stocking density?’’ At present, this question cannot be satisfactorily answered due to the lack of research in this area. Nevertheless, evidence suggests that the optimum stocking density in terms of minimal environmental degradation is likely to vary between environments, depending on factors such as soil texture, topography, and the presence/absence of subsurface drainage. It is also likely to vary over time, fluctuating both seasonally (taking into account seasonality of plant growth, soil moisture content) and annually (taking into account climatic fluctuations which influence plant growth and soil moisture). It is also likely to be scale‐dependent due to animal behavioral patterns and the tendency for animals to congregate in certain areas of the field. For example, a stocking density of 2 LSU per hectare may cause little damage in a 1‐ha field. However, the same stocking density in a 30 ha field may result in much higher levels of damage to the pasture because the livestock tend to concentrate at certain points in the field (drinking/feeding troughs, fence‐lines, gateways, sheltered spots) and therefore the eVective stocking density will be exaggerated at these points. There may be up to 60 LSU in less than 5% of the total area (in the above example). In the current market, the optimum stocking density for minimal environmental degradation may not be economically viable for many small conventional farmers. However, this is partly a problem associated with

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the value of produce and the amount that retailers and consumers are prepared to pay for environmentally and economically sustainable farming. At present, supermarket ‘‘price wars’’ on produce, such as milk and meat, threatens both the future of farming and the environment. Future research needs to quantify the link between damage to grassland soils/vegetation by grazing animals and the rate of soil erosion and sediment delivery to surface waters. Furthermore, there is a demand for scientists to determine the most sustainable stocking density for grasslands with diVering environmental characteristics. Clearly, this will need to consider the socioeconomic aspects of farming as well as the environmental consequences.

ACKNOWLEDGMENTS The authors are grateful to David Scholefield and Les Firbank of IGER North Wyke for comments on the chapter. We acknowledge UK Defra Projects PE0120 and PE0118 for funding that resulted in this work. IGER and SoilCIP are supported by core funding from the UK Biotechnology and Biological Sciences Research Council (BBSRC).

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Index A Abiotic stresses in rice genetic engineering and conventional breeding approaches for, 95 Acala SJ-2, radiation use efficiency (RUE) in, 213 A:Ci curves, in genetic variability, 216–7 Acid sulfate soils, genesis of, 11 Aerobic rice, for water-deficit areas, 110–1 Africa. See also Sub-Saharan Africa (SSA) drylands NERICA varieties, 90–4 rice-field weeds, 112 wild and cultivated rice species, 88–9 African rice gall midge (AfRGM), 56, 93–5 Africa Rice Center (WARDA), 78–9, 81, 84, 86, 90–7, 101, 114, 125 Ag/AgCl reference electrode, 13–5 Agriculture alternative and biological, 183 comprehensive management of environment, 195 Allophane, P adsorption on, 160, 169 Alternative wetting and drying (AWD), 110 Aluminum oxide, P adsorption on, 160, 169 Aluminum oxyhydroxides, P adsorption on, 160, 169 Ammonium oxalate extractable iron (Feox), 139 Amorphous Al(OH)3, P adsorption on, 141, 160, 169 Anaerobic dechlorination process, 8 Anaerobic soils, 5, 9–10 Anaerobiosis, 5–6 Animal species and age of grazing animals impact of defoliation, 255–6 influence on soil structural alteration, 241–2 Anopheles gambiae, 117 Arrhenius empirical rate law, 143 As and Fe K-edge EXAFS analyses, 157 Aspavia sp., 113 Atrazine, degradation of, 8–9

B Barbosalite (FeIIFeIII 2(PO4)2(OH)2), 153–4 Bas-fonds wetland, 66

Biodynamic agriculture, 183 Birnessite surfaces, P adsorption on, 141, 147 Boehmite (g-AlOOH), P adsorption on, 141, 148–9 Boliland wetland, 66, 77

C Ca2Ca3(PO4)3(OH, F), 137 Ca-kaolinite, 139–40 Calcite, 139 Carbonate-fluroapatite (Ca10(PO4, CO3)6F2–3), 140 Carbon-fiber electrodes, 16 Carbon hot spots, 5 Carbon metabolism, and photosynthesis, 209–29 13 C discrimination, in genetic variability, 217–9 Cecil clay, P adsorption on, 160 Chemical reaction, effect of temperature on, 143 Chemical transformation theory, 170 Chilo zacconius, 113 Chinese ecological agriculture (CEA) basic principles of, 188–91 case studies of, 195–202 characteristics of, 187–8 comprehensive management of agricultural environment, 195 development and achievements of, 191–3 ecological industry and, 199–202 education of farmers and, 203 energy exploitation, 194 integrated control techniques, 194 introducing new varieties, 194–5 lack of funds, 203 lack of market competitiveness, 204 lack of systematic theoretical research, 203 modern techniques application, 204–5 multilevel organic substance utilization, 193–4 objectives of, 188 optimization of, 205 poverty problem, 202–3 small production scale, 203

281

282

INDEX

Chinese ecological agriculture (CEA) (cont.) sustainable development strategy, 182 theory and practice, 202 vertically distributed farming, 193 Chinese Ecological Economics Society (CEES), 193 Chloroanilines acid, degradation of, 8–9 Chlorophyll fluorescence, 214–5, 226 Climate change, impact on rice intensification, 121–2 Coastal wetlands, 65–7 13 CO2 discrimination, in genetic variability, 218–9 CO2-exchange rates (CER), in photosynthesis, 210–4, 216 Comprehensive African Agricultural Development Program (CAADP), 125 Conservational farming, 183 Constant capacitance model (CCM), 145–7 Copiapite (MgFe4(OH)2(SO4)6–10H2O), 155 Coquimbite [(Fe,Al)2(SO4)3–9H2O], 155 Cotton C3 photosynthesis metabolism of, 216 Crandallite (CaAl3(PO4)2(OH)5-H2O), 140 Crop establishment for rice in SSA, 99 Crop straw, multilevel utilization of, 190 Cryptosporidium sp., 260, 265 Cultivation, relevance of redox potential (Eh) in, 6–7 Curuvularia sp., 113 Cyanobacteria, 136, 264

D Dambos wetland, 66 Dechlorinating bacteria, 8 Deepwater and mangrove rice ecosystems in SSA, 79–81 cropping systems, 80 cultivation practices and yields, 80–1 deepwater and mangrove swamp rice area, 80 production constraints, 81 Deepwater and mangrove swamps ecosystems in SSA, 72 Defoliation by grazing animals impact on grassland vegetation animal species and age, 255–6 stocking density, 256 vegetation response, 257–8

Dehalobacter restrictus, 8 Dehalococcoides ethenogenes, 8 Dehalospirillum multivorans, 8 Diffuse double layer theory, 145 Diffuse reflectance FTIR (DRIFTIR) spectroscopy, 148 1,1-Dimethylpiperidinium chloride. See Mepiquat chloride Diopsis macrophthalma, 113 ,0 -Dipyridyl dye, 34 Diseases and management in rice, in SSA, 113–14 DPL Acala 90 cotton, 218 Dryland rice ecosystems in SSA, 72–7 cropping systems, 75–6 cultivation practices and yields, 76 dryland rice area, 74–5 and nutrients management, 101 production constraints, 76–7 Dryland rice system, management technologies for, 119–20 Dryland soils in SSA, 67–8 Dyes, for assessing reduction in field, 34

E Earthworms (Lumbricina), 254 Echinochloa colona, 111, 113 Echinochloa crus-galli, 115 Ecological agriculture ecological industry and, 199–202 energy exploitation in, 194 international cooperation in, 205 pattern linked by methane, 201 township enterprise and, 205 in West, 183–4 Ecological industry, ecological agriculture and, 199–202 Ecosystem-based nutrients management, 101–5 Ecosystems of SSA, management technologies for rice in, 118–20 Electrophoretic mobility measurement studies for phosphate surface complexation on soil components, 147–8 Energy dispersive X-ray spectroscopy (EDX), 140 Entisols, 66–7, 139 Environmental degradation, by grazing animals, 239–40, 253, 266–72

INDEX Environmental issues, and rice intensification in SSA, 120–1 Environmental problems in China, 185–6 Eschericheria coli, 260, 265 Ethephon [(2-chloroethyl) phosphonic acid], 220 Eutrophication, 136, 240, 259, 263–5 Ex situ spectroscopic studies for soil components, 148–9

F Fadamas wetland, 66 Farmers education, and ecological agriculture, 203 Farming systems, in SSA, 58 Fe(III)-P salt [Fe4(PO4)3(OH)3], 153 Ferric ammonium sulfate [FeNH4(SO4)2-12H2O], 155 Ferrihydrite P adsorption on, 141–2, 148, 160, 162, 166–7 P desorption on, 161, 164 Fishponds, ecological agriculture pattern of, 194 Floating varieties rice, 90 Flowing tap water, electrodes testing in, 26

G G. barbadense, 218–9 leaf CER among, 211–2 G. davidsonii, rubisco activity in, 216 G. hirsutum, 218–9 leaf CER among, 211–2 rubisco activity in, 216 Gardia spp., 260, 265 Genetic engineering approaches incorporating resistance/tolerance, in rice, 95 Genetic variability A:Ci curves, 216–7 13 C discrimination, 217–9 chlorophyll fluorescence, 214–5 CO2 exchange, 210–4 in photosynthesis, 210–9 photosynthetic enzymes, 215–6 stomatal conductance, 210–4

283

Germplasm, breeding and variety development in SSA, 87–97 germplasm exchange by INGER-Africa, 96–7 glaberrima sativa crosses, 90–3 and irrigated rice varieties, 94–5 NERICA varieties for African drylands, 90–3 plant types and traditional rice varieties, 87–90 Glassy carbon electrodes, 16 G&P 74 þ cultivar, 217 Grain quality management, for rice in SSA, 114–5 Grasslands, 238–9 Grassland soils, impact of treading by grazing animals on, 240–1 Grassland vegetation impact of defoliation by grazing animals on, 254, 263 animal species and age, 255–6 stocking density, 256 vegetation response, 257–8 Grazing animals defoliation by on grassland vegetation, 254, 263 animal species and age, 255–6 stocking density, 256 vegetation response, 257–8 environmental degradation, 239–40, 253 land management strategy, 269–70 mitigation and damage reduction methods, 266 natural recovery of soil, 266–72 future research on, 272–3 soil structural alteration, 245–9 soil compaction, 242, 245–7, 263 soil poaching, 242, 247–51, 263 soil pugging, 242, 246–9, 263 treading impact on grassland soils, 240–1 natural recovery of soil, 266–72 Grazing animals excretion impact on grasslands livestock wastes as source of nutrients, 258–9 source of pathogens, 259–60 Grazing animals treading impacts of on soil structural alteration by soil fauna and, 254 soil hydrology and, 251–2 soil physical properties and, 249–51 vegetation growth and, 252–3

284

INDEX

Greenhouse gas, 9 Green Revolution in Asia, 85 in SSA, 124 Griphite (Fe3Mn2(PO4)2.5(OH)2) P adsorption on, 141, 147–8, 159–60, 166, 169 P desorption on, 161

H Harvested rice area, in SSA, 61 Heavy metals, relevance of redox potential (Eh) in, 7–8 Hematite P adsorption on, 141, 160 P desorption on, 161 High-effciency sustainable agriculture (HESA), 184 Histosols, 66–7 Hydrellia sp., 113 Hydric soils, 2, 44 Hydromorphic soil morphology, 3

I ICM for rice, 115–7 Inceptisols, 66–7 Industrial agriculture, 187 Information and communication technology (ICT) for rice, 122–3 Inland basins wetlands, 64–5 Inland valley toposequence-based nutrients management in, 105–6 wetlands, 65–6 Inositol phosphate (phytic acid), 137 Insect pests and biodiversity, techniques for, 194 and management, in rice in SSA, 113–4 In situ spectroscopic studies for phosphate surface complexation on soil components, 149–59 Integrated pest management (IPM) technologies, 113 Integrated plant nutrient systems (IPNS), 100 Integrated soil fertility management (ISFM), 100

Intensively managed grasslands impact of excretion by grazing animals on, 258–60 Intermittent irrigation technologies, 110 International Center for Tropical Agriculture (CIAT), 97 International cooperation, in ecological agriculture, 205 International Fertilizer Development Center (IFDC)-Africa, 100 International Institute for Tropical Agriculture (IITA), Nigeria, 91, 94, 97 International Rice Research Institute (IRRI), 86, 91, 94, 97, 122, 125 International Rice Testing Program (IRTP), 96 International sustainable agriculture, 184 Iron-coated (IRIS) tubes, for assessing reduction in field, 32–4, 44 Iron oxyhydroxides, P adsorption on, 160 Irrigated area, in SSA, 61 Irrigated rice ecosystem based, nutrients management, 102 Irrigated rice farming systems and water use, 108–11 Irrigated wetland rice crops and nutrients, 100 Irrigated wetland rice ecosystems in SSA, 72, 81–5 cropping systems, 83 cultivation practices and yields, 83–4 in humid forest and savanna zones, 82 irrigated wetland rice area, 83 production constraints, 84–5 in Sahel, 82 in tropical highlands, 82–3 Ischaemum rugosum, 115

J Jarosite (KFe3(SO4)2(OH)6), 155 Jurisdictional wetland, 3

K Kaolinite, P adsorption on, 160

INDEX L Land management strategy, 269–70 Langmuir equation, 138–9 Late-duration erect varieties rice, 90 Lepidocrocite, P adsorption on, 166 Linear combination (LC) XANES fit analyses, 152 Liuminying ecological farm, 200–1 Livestock management and grazing animals, 266–9 Livestock units (LSU) per hectare, 243, 272 Livestock wastes as source nutrients of, 258–9 pathogens of, 259–60 Lolium perenne, 259 Low-energy agriculture, 183 Low-input sustainable agriculture (LISA), 184

M Maliarpha sp., 113 Management technologies for rice, in ecosystems of SSA, 118–20 Mangrove swamps rice system management technologies, 119–20 Marigots wetland, 66 Market competitiveness, and ecological agriculture, 204 MD 65–11, leaf CER among, 212 Mepex, plant growth regulator, 219 Mepichlor, plant growth regulator, 219 Mepiquat chloride, plant growth regulator, 219–20 Metal oxides, phosphate adsorption on, 141–2 Metal-oxyhydroxide interface, P adsorption mechanisms, 152 Methane (CH4) ecological agriculture pattern linked by, 201 relevance of redox potential in, 9–10 Methanotrophic bacteria, 9 Metolachlor, degradation of, 8–9 Modern techniques application, in ecological agriculture, 204–5 Mossbauer spectroscopy, 139 Mountainous regions, ecological agriculture in, 195–7

285

Mountain rice, 90 Mulberry trees, ecological agriculture pattern of, 194 Multilevel organic substance and ecological agriculture, 193–4 Multilevel utilization, of crop straw, 190

N Natural agriculture, 183 NERICA varieties, 57–8 for African drylands, 90–3 for African rainfed wetlands, 93–4 breeding of, 93–4 traits of dryland, 91–3 Nernst equation, 29 New Partnership for African Development (NEPAD), 125 New Zealand loamy soils, P adsorption on, 145, 160 Nucleic acid, 137 Nutrients and irrigated wetland and rainfed dryland rice crops, 100 livestock wastes as source of, 258–9 for rice research and technology development in SSA, 99–107 Nymphula depunctalis, 113

O Okaihau gravelly clay, P adsorption on, 160 Oligotrophic-mesotrophic waters, 264 Organic agriculture, 183 food, 205 substance utilization, in CEA, 193–4 Orsylia oryzivora, 113 Oryza barthii, 88 Oryza brachyantha, 88 Oryza eichingeri, 89 Oryza glaberrima, 57, 87, 89–93, 112, 120, 124 Oryza longistaminata, 88 Oryza punctata, 88 Oryza rufipogon, 108 Oryza sativa, 57, 87, 89–93, 108, 112, 124 Oryza stapfii, 89

286

INDEX

Oswald-step rule, 170 Over compensatory growth, 256 Oxalate extractable P (Pox), 140 Oxisols, 139

P Paints of the earth. See Redoximorphic features of soil Panicum laxum, 111 Paspalum scrobiculatum, 113 Pathogenic contamination by grazing animals, 265 Pathogens, livestock wastes as source of, 259–60 P chemistry. See Phosphorus (P) chemistry Pentachlorophenol, degradation of, 8–9 Perennial ryegrass (Lolium perenne), 245 Pesticides, redox potential (Eh), 8–9 Petrochemical-based agriculture, 182 PGR-IV, plant growth regulator, 220 PH effects, 140–1 Phosphate adsorption on empirical approaches, 138–9 I effects on P surface complexation, 145–6 kinetics, 142, 144, 162–3 metal oxides, 141–2 pH effects on, 140–1 phyllosilicate minerals, 142–3 physicochemical properties of soils and, 139–40 soil components, 149–59 soils, 138–9 temperature effects, 143–4 Phosphate desorption in soil components higher energy binding through chemical reconfiguration, 170–1 kinetics, 164 residence time effects theory, 159–60 slow desorption process and hysteresis, 161–5 surface precipitation, 167–70 Phosphate surface complexation on soil components, 146–59 electrophoretic mobility measurement studies, 147–8 modeling approaches, 146–7 spectroscopic studies, 148–59 Phosphoenolpyruvate carboxylase, 216

Phosphorus (P) chemistry, 136–8 Photoinhibition, 228 Photosynthesis carbon metabolism and, 209–29 down regulation of, 228 genetic variability A:Ci curves, 216–7, 224 13 C discrimination, 217–9 chlorophyll fluorescence, 214–5 CO2 exchange, 210–4 photosynthetic enzymes, 215–6 stomatal conductance, 210–4 management and environmental effects moisture stress, 223–5 plant growth regulators, 219–20 plant nutrition, 221–2 soil fertility, 221–2 temperature, 225–8 nonstomatal reduction in, 224 process, 209, 229 Photosynthetic enzymes, in genetic variability, 215–6 Phyllosilicate minerals, phosphate adsorption on, 142–3 Phyllosilicate oxyhydroxides, P adsorption on, 160 Pix, plant growth regulator, 220 P K-edge XANES techniques, 151–4, 157 Plant diseases controlling, integrated techniques for, 194 Plant nutrients, 209 and vitality, relevance of redox potential (Eh) in, 7 Plant types and traditional rice varieties, in SSA, 87–90 Pluvial rice, 72 Pollution-free food, 205 Potentiometric measuring techniques data-recording unit, 24 limit of measurement, 20–1 problems in, 20–4 for redox potential (Eh), 11–25 reference cells, 13–5 technical problems, 21–4 testing electrodes prior to installation, 24–5 transmission lines, 23–4 working redox electrodes, 15–23 Poverty, and ecological agriculture, 202–3 P surface complexation, I effects on, 145–6 Pyricularia oryzae, 113

INDEX Q Quadruple layer model theory, 145

R Rainfed dryland rice crops, nutrients removal, 100 Rainfed rice farming, managing water for, 107–8 Rainfed wetland rice ecosystem, 77–9 based nutrients management, 101–2 cropping systems, 78 cultivation practices and yields, 78–9 management technologies for, 119–20 production constraints, 79 rainfed wetland rice area, 78 in SSA, 72 varieties rice, 90 Rainwater harvesting (RWH), 76, 107 Redox environments, with electron acceptors, 31 Redox field data interpretation correction of field data, 25–7 for pH, 26–7 to standard hydrogen electrode, 25–6 pooling of long-term data sets, 30–2 variability, 27–30 short-term changes due to soil chemistry, 30 spatial variability, 27–9 temporal variability, 29–30 Redoximorphic features of soil, 3, 11 Redox potential (Eh) alternative methods for assessing in field, 32–6 definition of, 4–5 interpreting of, 43–4 irreversibility, 20 of marine sediments, 4 measurement data-recording unit, 24 field installation and procedure for, 36–43 limit of, 20–1 problems in, 20–4 procedure for, 41 reference cell, 13–5, 21 techniques, 11–25

287

testing electrodes prior to installation, 24–5 working redox electrodes, 15–23 mixed potentials, 21 potentiometric measuring techniques, 11–25 relevance in soil science, 5–11 agriculture, 6–7 cultivation, 6–7 heavy metals, 7–8 methane, 9–10 pesticides, 8–9 plant nutrients and vitality, 7 radionuclides, 9 soil genesis, 10–1 toxic organics, 8 slow reaction kinetics, 20 variation in for soil, 37–8 wetland environmental issues and, 2–5 of ZoBell’s solution as function of temperature, 19 Redox potential (Eh) measurement data-recording unit, 24 field installation and procedure, 36–4 common field problem, 41–2 field pH measurements, 42–3 interpreting Eh, 43–4 number of electrodes to install, 36–8 process used to install electrodes, 40 Pt electrode installation, 38–40 reading electrodes, 41 voltage drift problem, 41–2 limit of, 20–1 problems in, 20–4 procedure for, 41 reference cell, 13–5, 21 techniques, 11–25 testing electrodes prior to installation, 24–5 working redox electrodes, 15–23 Redox process, 3–4 Reduction in fields, alternative methods for assessment dyes, 34 iron-coated tubes, 32–4 zero valence iron rods, 35–6 Reference cells, for potentiometric measuring techniques, 13–5, 21 Reference electrodes, standard half-cell potentials of, 27

288

INDEX

Regenerative agriculture, 183 Residence time effects on phosphate adsorption and desorption in soils higher energy binding, 170–1 inter and intraparticle diffusion, 165–7 residence time effects theory, 159–60 slow adsorption, 160–1 slow desorption process and hysteresis, 161–5 solid-state diffusion, 165–7 surface precipitation, 167–70 Residence time effects theory, for phosphate desorption in soils, 159–60 Rhyncosporium oryzae, 113 Rice areas distribution in SSA, 62 and yield trends, in SSA, 62–3 Rice cropping, essential nutrients in, 103–4 Rice cultivation, in SSA, 57 Rice demand and supply, in SSA, 58–63 Rice ecosystems, in SSA dryland rice ecosystems, 72–7 cropping systems, 75–6 cultivation practices and yields, 76 dryland rice area, 74–5 production constraints, 76–7 related environmental and human disease problems in, 73–4 by surface-water regimes, 72 wetland rice ecosystems, 77–85 deepwater and mangrove rice ecosystems, 79–81 irrigated wetland rice ecosystems, 81–5 rainfed wetland rice ecosystem, 77–9 Rice farming systems, diversification of in SSA, 115 Rice intensification in SSA environmental issues related to, 120–1 biodiversity, 120 land degradation, 121 pollution, 121 production, 120 water, 121 impact of climate change, 121–2 information and communication technology (ICT) based system, 122–3 policy support for, 123–4

technology delivery and deployment issues, 122–3 vector-borne human diseases and, 117–20 Rice production constraints, 56–7, 85–6, 125 human resource constraints, 85–6 lack of education, 85 lack of researchers, 85 physical, biological, and management constraints, 85 socioeconomic and policy constraints, 86 in SSA, 85–6 in SSA area and yield trends, 62–3 constraints to, 56–7 and imports in, 59–60 and yield of unmilled rice in, 61 Rice research and technological development in SSA from breeding to milling, 114–5 crop establishment, 99 diseases and their management, 113–4 diversification of rice farming systems, 115 germplasm, breeding and variety development, 87–97 germplasm exchange by INGER-Africa, 96–7 glaberrima sativa crosses, 90–3 improved irrigated rice varieties, 94–5 NERICA varieties, 90–3 plant types and traditional rice varieties, 87–90 grain quality management, 114–5 ICM for rice, 115–7 insect pests and management, 113–4 nutrients management, 99–107 and dryland rice ecosystem, 101 ecosystem-based, 101–5 and irrigated rice ecosystem, 102 and rainfed wetland ecosystem, 101–2 site-specific nutrient management, 106–7 toposequence-based in Inland valley, 105–6 seed production and distribution services, 97–8 and seed sector services and seed research status, 97–8 formal seed supply systems, 97–8

INDEX informal seed supply systems, 98 national seed laws and variety release process, 98 water management for rainfed and irrigated areas, 107–111 weeds management, 111–3 Rice soil resources in SSA dryland soils, 67–8 wetland soils, 68–71 Rice yellow mottle virus (RYMV), 57, 79, 93–4, 96 River floodplains wetlands, 65–6 Rockbridgeite ((FeII,MnII)FeIII4 (PO4)3(OH)5), 153–4 Rockefeller Foundation, 125 Rotavirus sp., 260, 265 Rottboellia exaltata, 113 Rribulose-1,5-bisphosphate carboxylaseoxygenase (Rubisco), 215 RWH. See Rainwater harvesting (RWH)

S Salinity management for irrigation, 109 Salmonella sp., 260, 265 Sarocladium sp., 113 Scanning transmission electron microscopy (STEM), 140 Schistosoma haematobium, 117 Schistosoma mansoni, 117 Scirpophaga sp., 113 Season erect varieties rice, 90 Seed production and distribution services in SSA, 97–8 existing seed sector services and seed research status, 97–8 formal seed supply systems, 97–8 informal seed supply systems, 98 national seed laws and variety release process, 98 Self-contained agricultural ecosystem, 183 Semiconductors as electrode sensors, 16 Sesamia calamistis, 113 Setaria pumila, 113 Shifting cultivation, 75 Site-specific nutrient management, 100, 106–7 Slash-and-burn cultivation, 75 Small production scale, in ecological agriculture, 203

289

Soil, morphological and redoximorphic features of, 3 Soil compaction, 242, 245–7, 263 Soil components oxyanion adsorption on, 141 phosphate adsorption on, 138–46 empirical approaches, 138–9 higher energy binding through chemical reconfiguration, 170–1 I effects on P surface complexation, 145–6 metal oxides, 141–2 pH effects on, 140–1 phyllosilicate minerals, 142–3 physicochemical properties of soils and, 139–40 residence time effects theory, 159–60 slow adsorption, 160–1 solid-state inter and intraparticle diffusion, 165–7 surface precipitation, 167–70 temperature effects, 143–4 phosphate desorption in higher energy binding, 170–1 inter and intraparticle diffusion, 165–7 residence time effects theory, 159–60 slow desorption process and hysteresis, 161–5 solid-state diffusion, 165–7 surface precipitation, 167–70 phosphate surface complexation on, 146–59 electrophoretic mobility measurement studies, 147–8 ex situ spectroscopic studies, 148–9 in situ spectroscopic studies, 149–59 modeling approaches, 146–7 physicochemical properties affect on phosphate retention, 139–40 Soil damage, and stocking density, 242 Soil deformation and treading, 244 Soil erosion problems control technology, 195 and grazing animals, 262–3 Soil fauna and treading, 254 Soil genesis and redox potential, 10–11 Soil hydrology and treading, 251–2 Soil moisture, influence on soil structural alteration, 243–4

290

INDEX

Soil physical condition, 266–72 Soil physical properties and treading, 249–51 Soil-plant-animal system, phosphorus in, 136 Soil poaching, 242, 247–51, 263 Soil pugging, 242, 246–7, 263 Soil redox potential (Eh) measurement data–recording unit, 24 limit of measurements, 20–1 problems in, 20–4 reference cells, 21 technical problems, 21–4 techniques, 11–25 testing electrodes prior to installation, 24–5 transmission lines, 23–4 working electrodes, 21–3 Soil sedimentation problems and grazing animals, 240, 262–3 Soil structural alteration factors influencing amount and form, 241–5 animal species and age, 241–2 soil moisture, 243–4 stocking density, 242–3 vegetation cover, 244–5 forms resulting from treading by grazing animals, 245–9 soil compaction, 242, 245–7, 263 soil poaching, 242, 247–51, 263 soil pugging, 242, 246–9, 263 impacts of grazing animals treading, 249–54 soil fauna and, 254 soil hydrology and, 251–2 soil physical properties and, 249–51 Spermacoce ruelliae, 113 S-6 Pima cotton, 218 S-7 Pima cotton, 218 SSNM. See Site-specific nutrient management State Bureau of Environmental Protection, China, 191 State Science and Technology Commission, China, 191 Stenocoris claviformis, 113 Stocking density of grazing animals by defoliation, 256, 263 and soil structural alteration, 242–3 Stomatal conductance in genetic variability, 210–4 in photosynthesis, 210–4

Strengite (FeIII(PO4)–2H2O), 153–4, 169 Striga sp., 111 Sub-Saharan Africa (SSA) agroclimatic zones, 71–85 and high-yielding rice varieties, 124–5 countries with large areas in rice ecologies/ecosystems, 75 distribution of rice areas in, 62 ecosystems, 71–85 environmental and human disease problems in, 73 farming systems in, 58 green revolution in, 124 harvested rice area, 61 hunger and malnutrition in, 56–7 irrigated area, 61 per capita rice consumption in, 58–9 plant types and traditional rice varieties of, 87–90 poverty and food insecurity in, 56–7 prevalence of diseases, 57 rice area and yield trends in, 62–3 rice cultivation in, 57 rice demand and supply in, 58–63 rice intensification in environmental issues, 120–1 impact of climate change, 121–2 information and communication technology (ICT) system, 122–3 policy support for, 123–4 technology delivery and deployment issues, 122–3 vector-borne human diseases and, 117–20 rice production area and yield trends, 62–3 challenges to enhance, 124–5 constraints to, 56–7, 85–6, 125 and imports in, 59–60 and yield of unmilled rice in, 61 rice research and technology development in, 86–117 rice soil resources, 63–7 wetlands, 63–7 Surface complex models (SCM) approaches to phosphate surface complexation on soil components, 146–7 Surface-water regimes, rice ecosystems in SSA by, 72

INDEX Surface waters in intensively managed grasslands impact of excretion by grazing animals on, 258–61 impact of grazing animals on quality of, 239–40, 260–5 eutrophication, 240, 263–5 pathogenic contamination, 265 soil erosion and sedimentation problems, 240, 262–3 impacts of treading and defoliation by grazing animals, 261 Sustainable agriculture and rural development (SARD), 184 Sustainable agriculture (SA) interactions in, 187 research on, 184, 186–7 Sustainable development strategy, for ecological agriculture, 182

T TAMCOT HQ95 cultivar, 217 Technology delivery and deployment, for rice intensification in SSA, 122–3 Temperature effects, on phosphate adsorption, 143–4 Tetrachloroethene (PCE), reductive dechlorination of, 8 Thanetophorus cucumeris, 113 Tidal rice, 81 Tinticite (Fe4(PO4)4(OH)6–7H2O), 169 Toposequence-based nutrients management in Inland valley, 105–6 Township enterprise, and ecological agriculture, 205 Toxic organics, relevance of redox potential (Eh) in, 8 Transmission lines, problems in measurement of Eh due to, 23–4 Transplanted rice system, management technologies for, 119–20 Treading by grazing animals impact on grassland soils, 240–1 soil structural alteration forms resulting from, 245–9 soil compaction, 242, 245–7, 263 soil poaching, 242, 247–51, 263 soil pugging, 242, 246–9, 263 Trianthema portulacastrum, 113

291

Trichloroethene (TCE), reductive dechlorination of, 8 2,4,5-Trichlorophenoxyacetic acid, degradation of, 8–9 Trifolium repens, 259 Tube-well irrigation systems, 109

U United Nations Environment Program (UNEP), 192 Upland rice, 72 USAID, 125

V Vant Hoff’s law, 29 Variable charge minerals, pH effects on phosphate adsorption on, 140–1 Vector-borne human diseases in SSA bilharzia incidence, 117 malaria incidence, 117 and rice intensification issues, 117–20 schistosomiasis incidence, 117 Vegetation cover, influence on soil structural alteration, 244–5 Vegetation growth and treading, 252–3 Vegetation in intensively managed grasslands, 258–60 Vegetation response of grazing animals, 256 Vertically distributed farming, in ecological agriculture, 193 Vivianite (Fe(II)3(PO4)2), 169

W Waste management and grazing animals, 270–2 Waste water for irrigation, 109 Water-collecting ecological agriculture in Western China, 197–9 Water productivity practices in irrigated rice farming, 110 Water quality of surface waters in intensively managed grasslands impact of grazing animals, 260–5 eutrophication, 263–5

292

INDEX

Water quality of surface waters in intensively managed grasslands (cont.) pathogenic contamination, 265 soil erosion and sedimentation problems, 262–3 impacts of treading, defoliation, and excretion by grazing animals, 261 Water-saving technologies, 110 Wavellite (Al3(PO4)2(OH)3–5H2O), 137, 140 Weeds management for rice in SSA, 111–3 Well-poised redox buffer solution, electrodes testing in, 26 West, ecological agriculture in, 183–4, 188 West Africa, agroecological zones of, 68 Western China, water–collecting ecological agriculture in, 197–9 Wetland NERICAs, 94 Wetland rice ecosystems in SSA, 77–85 deepwater and mangrove rice ecosystems, 79–81 irrigated wetland rice ecosystems, 81–5 rainfed wetland rice ecosystem, 77–9 Wetlands areas, 64–5 characteristics of, 68–71 definition, 64 distribution of, 64–5 potential source of rice production in SSA, 63–7 rice soils, 70–1 in anthropic (artificial) wetlands on terraced landscapes, 70 classification, 70–1 converted from natural wetlands, 70 potential for rice production in Africa, 70–1 problem soils, 70–1 types and characteristics of, 64–7 coastal wetlands, 65–7 inland basins, 64–5 inland valleys, 65–6 river floodplains, 65–6 Wetland soils, 3 White clover (Trifolium repens), 245

Wild and cultivated rice species, in Africa, 88–9 Working redox electrode buffer or poise test solutions for calibration of, 19 carbon-fiber electrodes, 16 design and construction, 17–8 glassy carbon electrodes, 16 graphite electrodes, 16 metals as electrode sensors, 15–6 for potentiometric measuring techniques, 15–23 pretreatments, 18 problems in measurement of Eh, 21–3 absorptive contamination, 21, 23 insufficient galvanic contact and polarization of electrodes, 23 semiconductors as electrode sensors, 16 testing procedure, 18–20 Wulian County benefits of ecological agriculture in, 197 ecological agriculture construction in, 196–7

X Xanthomonas campestris pv. oryzae, 113 X-ray absorption near edge structure spectroscopy (XANES), 151–2 X-ray absorption spectroscopy (XAS), 138, 149 X-ray photoelectron spectroscopy (XPS), 148

Z Zaleya pentandra, 113 Zero valence iron rods, for assessing reduction in field, 35–6 ZoBell’s solution, redox potential as function of temperature, 19