Hypoxia in the Northern Gulf of Mexico

An Update by the EPA Science Advisory Board
 

 United States 

Environmental Protection 

Agency 

 EPA Science Advisory Board 

 1400F 

 Washington, D.C. 

 EPA-SAB-08-003 

 December 2007 

 www.epa.gov/sab 

The EPA Science Advisory Board (SAB) of the U.S. Environmental Protection Agency is a body 

of external scientists who provide independent advice to the EPA Administrator on scientific and 

technical issues. The SAB was established in its present form by the Congress in 1978. The 

SAB's principal mission is to review the quality and relevance of the scientific information being 

used to support Agency decisions, programs and strategies and provide broad strategic advice 

on scientific and technological issues. In addition, the SAB occasionally conducts self-initiated 

studies or special studies at the request of the Administrator to address current and future 

environmental issues. The SAB meets in public sessions, and its committees and review panels 

are designed to include a diverse and technically balanced range of views, as required by the 

Federal Advisory Committee Act. 

Cover photo: Courtesy of Louisiana State University's Earth Scan Laboratory. Satellite image of 

the Louisiana coast and Northern Gulf of Mexico captured by Oceansat-1 using Ocean Color 

Monitor, April 12, 2007. This image is a true color enhancement using the red, green and blue 

channels of the electromagnetic spectrum. The image shows high levels of suspended 

sediments delivered to the Gulf via the Mississippi and Atchafalaya Rivers. 
 UNITED STATES ENVIRONMENTAL PROTECTION AGENCY 

 WASHINGTON D.C. 20460 

 

OFFICE OF THE ADMINISTRATOR 

EPA SCIENCE ADVISORY BOARD 

December 21, 2007 

EPA-SAB-08-003 

Honorable Stephen L. Johnson 

Administrator 

U.S. Environmental Protection Agency 

1200 Pennsylvania Avenue, N.W. 

Washington, D.C. 20460 

 Subject: Hypoxia in the Northern Gulf of Mexico: An Update 

 by the EPA Science Advisory Board 

Dear Administrator Johnson: 

 Over a year ago, the EPA Office of Water (OW) asked the Science Advisory 

Board (SAB) to evaluate the most recent science on the hypoxic zone in the Gulf of 

Mexico as well as potential options for reducing the size of the zone. The hypoxic zone, 

an area of low dissolved oxygen that cannot support most marine life, has been 

documented in the Gulf of Mexico since 1985 and was most recently measured at 20,500 

km2. The SAB was asked to address the science that has emerged since the 2000 

publication of An Integrated Assessment: Hypoxia in the Northern Gulf of Mexico 

(Integrated Assessment), the seminal study by the Committee on Environment and 

Natural Resources that served as the basis for activities coordinated by the Mississippi 

River/Gulf of Mexico Watershed Nutrient Task Force. The SAB was also asked to 

address the most recent science on water quality in the Mississippi Atchafalaya River 

Basin, an area of 31 States and Tribes that drains approximately 40% of the contiguous 

United States. Further, the SAB was asked to discuss options for reducing hypoxia in 

terms of cost, feasibility and social welfare. To address this question, the SAB found it 

necessary to discuss recent research on water quality as well as research on policy 

options, in particular, those policies that create economic incentives. 

 

 Following OW’s request, the Science Advisory Board Staff Office convened an 

expert panel under the auspices of the chartered SAB. This SAB Panel consists of 21 

distinguished scientists from academia, industry and government agencies with expertise 

in the fields of oceanography, ecology, agronomy, agricultural engineering, economics 
ii

and other fields. Over the past year, the SAB Panel held numerous public meetings and 

considered information from invited speakers as well as over 60 sets of public comments 

in the development of this report. 

 

 In issuing the attached report, the SAB reaffirms the major finding of the 

Integrated Assessment, namely that contemporary changes in the hypoxic area in the 

northern Gulf of Mexico are primarily related to nutrient loads from the Mississippi 

Atchafalaya River basin. If the size of the hypoxic zone is to be reduced, the SAB finds 

that a dual nutrient strategy is needed that achieves at least a 45% reduction in both 

riverine total nitrogen flux and riverine total phosphorus flux. The SAB offers these as 

initial targets while stressing the importance of moving in a directionally correct fashion 

and adjusting policy adaptively on the basis of future data, changing conditions and 

lessons learned. Climate change will likely contribute to changing conditions. A number 

of studies have suggested that climate change will create conditions where larger nutrient 

reductions, e.g., 50 – 60% for nitrogen, would be required to reduce the size of the 

hypoxic zone. An adaptive management approach, coupling nutrient reductions with 

continuous monitoring and evaluation, can provide valuable lessons to improve future 

decisions. 

 The SAB was asked to comment on the Task Force’s goal of reducing the size of 

the hypoxic zone to 5,000 km2

 by 2015. With respect to the time frame, the SAB finds 

that such a significant reduction is not likely to be achievable over the next eight years. 

We conclude this for two reasons. First, there is limited current movement to implement 

policies, programs and strategies that reduce nutrients. Second, there are time lags 

between reductions in nutrient inputs and the response of the ecological system. Hence, 

while the 5,000 km2

 target remains a reasonable objective in an adaptive management 

context; it may no longer be possible to achieve this goal by 2015. This makes it even 

more important to proceed in a directionally correct fashion to manage factors affecting 

hypoxia than to wait for greater precision in setting the goal for the size of the zone. 

 

 The SAB underscores that in considering management strategies to reduce Gulf 

hypoxia, EPA should consider the many benefits of nutrient reduction in the Mississippi 

Atchafalaya River basin. Such “co-benefits” include improved groundwater and surface 

water quality, wildlife and biodiversity, recreation, soil quality, greenhouse gas reduction 

and carbon sequestration. In many cases, co-benefits may exceed the benefits of hypoxia 

reduction. 

 

 Finally, to reduce hypoxia in the Gulf, a systems view, looking at all sources and 

effects, is needed. The SAB urges the Agency to consider its options with respect to both 

non-point and point sources. Non-point sources have long been acknowledged as the 

primary source of nutrient loadings, however the SAB finds point sources are a more 

significant contributor than previously thought. Atmospheric deposition of nitrogen is 

also playing a role in hypoxia. In addition, it may be necessary to confront the conflicts 

between hypoxia reduction as a goal on the one hand and incentives provided by current 

agricultural and energy policy on the other. Some aspects of current agricultural and 

energy policies are providing incentives that contribute to greater nutrient loads now and 
iii

in the future. The SAB recognizes that if agricultural, environmental, and energy policies 

are to be aligned to support hypoxia reduction, cooperation across a broad spectrum of 

interests, including the highest levels of government, would be required. We note that 

regulatory options under the Clean Water Act, an area within EPA’s purview, are 

addressed by the National Academy of Sciences (NAS) in its recent study, the 

“Mississippi River and the Clean Water Act.” As pointed out by the NAS, EPA has 

regulatory authority under the Clean Water Act to address watershed wide issues. 

 

 The Executive Summary in the attached Advisory highlights the SAB’s findings 

and recommendations with more detailed science presented in the main body of the 

report. We appreciate the opportunity to provide advice on this important and timely 

topic and look forward to receiving your response. 

 Sincerely, 

 

/Signed/ /Signed/ 

Dr. M. Granger Morgan Dr. Virginia Dale 

Chair, Science Advisory Board Chair, SAB Hypoxia Advisory Panel 
iv

NOTICE 

 This report has been written as part of the activities of the EPA Science Advisory 

Board, a public advisory committee providing extramural scientific information and 

advice to the Administrator and other officials of the Environmental Protection Agency. 

The SAB is structured to provide balanced, expert assessment of scientific matters related 

to problems facing the Agency. This report has not been reviewed for approval by the 

Agency and, hence, the contents of this report do not necessarily represent the views and 

policies of the Environmental Protection Agency, nor of other agencies in the Executive 

Branch of the federal government. Mention of trade names or commercial products do 

not constitute a recommendation for use. Reports of the EPA SAB are posted at: 

http://www.epa.gov/sab. 
v

ACKNOWLEDGEMENTS 

 EPA’s Science Advisory Board Hypoxia Advisory Panel would like to 

acknowledge many individuals who provided their scientific perspectives for the Panel’s 

consideration in the development of this report. 

 Invited Speakers: 

 James Ammerman, Rutgers University, Nutrient Dynamics

 Rich Alexander, U.S. Geological Survey, SPARROW Model

 Jeff Arnold, U.S. Department of Agriculture, SWAT Model

 James Baker and Dean Lemke, UMRSHNC, Upper Mississippi Symposia 

Summary

 Robert Dean, University of Florida, Drawing Louisiana’s New Map

 Steven DiMarco, Texas A&M University, Physical Oceanography in the 

Gulf

 Katie Flahive, U.S. Environmental Protection Agency, Status of the 

Management Actions Reassessment Team (MART) Report

 Rick Greene (EPA) and Alan Lewitus (National Oceanic and Atmospheric 

Administration), Gulf Science Symposia Summary

 Dan Jaynes, U.S. Department of Agriculture, Agricultural N & P 

Management Approaches 

 Bob Kellogg, U.S. Department of Agriculture, Status of the Conservation 

Effectiveness Assessment Program (CEAP)

 Tim Miller, U.S. Geological Survey, Monitoring Activities in the 

Mississippi River basin

 Marc Ribaudo, U.S. Department of Agriculture, Costs and Benefits of 

Methods to Reduce Nutrient Loads 

 Don Scavia, University of Michigan, 1) Science and Policy Context & 2) 

Hypoxia Forecast Models 

 Janice Ward, U.S. Geological Survey, Fate and Transport Symposia 

Summary

Invited Technical Reviewers: 

• Mark Alley, Virginia Tech 

• Walter Dodds, Kansas State University 

• Madhu Khanna, University of Illinois 

• William Wiseman, Jr., National Science Foundation 

Public Commenters: 

• James Baker, Iowa Department of Agriculture and Land Stewardship 

• Victor Bierman, Donald Boesch, John Day, Robert Diaz, Dubravko Justic, 

Dennis Keeney, William Mitsch, Nancy Rabalais, Gyles Randall, Donald 

Scavia, and Eugene Turner, Contributors to the Integrated Assessment 
vi

• Donald Boesch, University of Maryland Center for Environmental Science 

• Darrell Brown, EPA Office of Water 

• Daniel Coleman, O’Brien & Gere 

• Richard Cruse, Iowa State University 

• Doug Daigle, Lower Mississippi River Sub-basin Committee on Hypoxia 

• Bob Diaz, Virginia Institute of Marine Sciences 

• Michael Duffy, Iowa State University 

• Nancy Erickson, Illinois Farm Bureau 

• Jason Flickner, Kentucky Waterways Alliance 

• Norman Fousey, U.S. Department of Agriculture 

• James Fouss, U.S. Department of Agriculture 

• Doug Gronau, Iowa Farm Bureau Federation 

• Ben Grumbles, Assistant Administrator for EPA’s Office of Water 

• Stephen Harper, O’Brien & Gere 

• Chuck Hartke, Illinois Department of Agriculture 

• Susan Heathcote, Iowa Environmental Council 

• Matthew Helmers, Iowa State University 

• Ed Hopkins, Sierra Club 

• Chris Hornback, National Association of Clean Water Agencies 

• Illinois Department of Agriculture 

• Thomas Isenhart, Iowa State University 

• Dan Jaynes, U.S. Department of Agriculture 

• Doug Karlen, U.S. Department of Agriculture 

• Dennis Keeney, Institute for Agriculture and Trade Policy 

• Louis Kollias, Metropolitan Water Reclamation District of Greater 

Chicago 

• Dean Lemke, Iowa Department of Agriculture and Land Stewardship 

• Alan Lewitus and David Kidwell, National Oceanic and Atmospheric 

Administration 

• Antonio Mallarino, Iowa State University 

• Mark Maslyn, American Farm Bureau Federation 

• Dennis McKenna, Illinois Department of Agriculture 

• Mississippi River Water Quality Cooperative (MSWQC) 

• Bill Northey, Iowa Secretary of Agriculture 

• Don Parrish, American Farm Bureau 

• Paul Patterson, City of Memphis 

• Jean Payne, Illinois Fertilizer and Chemical Association 

• Michelle Perez, Environmental Working Group 

• Bob and Kristen Perry, Missouri Clean Water Commission 

• Nancy Rabalais, Louisiana Universities Marine Consortium 

• Russell Rasmussen, Wisconsin Department of Natural Resources 

• Jack Riessen, Iowa Department of Natural Resources 

• Rick Robinson, Iowa Farm Bureau 

• Matt Rota, Gulf Restoration Network 
vii

• John Sawyer, Iowa State University 

• Al Schafbuch, Affiliation not identified 

• Theodore Slawecki, Limno-Tech, Inc. 

• Tim Strickland, U.S. Department of Agriculture 

• Richard Swenson, U.S. Department of Agriculture 

• Michael Tate, Kansas Department of Health and Environment 

• Steve Taylor, Environmental Resource Coalition 

• Mark Tomer, U.S. Department of Agriculture 

• Eugene Turner, Louisiana State University 

• Ford B. West, The Fertilizer Institute 

• Wendy Wintersteen, Iowa State University 
viii

Science Advisory Board 

Hypoxia Advisory Panel 

U.S. Environmental Protection Agency 

CHAIR 

Dr. Virginia Dale, Corporate Fellow, Environmental Sciences Division, Oak Ridge 

National Laboratory, Oak Ridge, TN 

MEMBERS 

Dr. Thomas Bianchi, Professor, Oceanography, Geosciences, Texas A&M University, 

College Station, TX 

Dr. Alan Blumberg, Professor, Civil, Environmental and Ocean Engineering, Stevens 

Institute of Technology, Hoboken, NJ 

Dr. Walter Boynton, Professor, Chesapeake Biological Laboratory, Center for 

Environmental Science, University of Maryland, Solomons, MD 

Dr. Daniel Joseph Conley, Professor, Marie Curie Chair, GeoBiosphere 

Centre, Department of Geology, Lund University, Lund, Sweden 

Dr. William Crumpton, Associate Professor & Coordinator of Environmental Programs, 

Ecology, Evolution, and Organismal Biology, Iowa State University, Ames, IA 

Dr. Mark David, Professor, Natural Resources & Environmental Sciences, University of 

Illinois, Urbana, IL 

Dr. Denis Gilbert, Research Scientist, Ocean and Environment Science Branch, 

Maurice-Lamontagne Institute, Department of Fisheries and Oceans Canada, Mont-Joli, 

Quebec, Canada 

Dr. Robert W. Howarth, David R. Atkinson Professor, Department of Ecology and 

Evolutionary Biology, Cornell University, Ithaca, NY 

Dr. Catherine Kling, Professor, Department of Economics, Iowa State University, 

Ames, IA 

Dr. Richard Lowrance, Research Ecologist, Southeast Watershed, Agricultural 

Research Service, USDA, Tifton, GA 

Dr. Kyle Mankin, Associate Professor, Biological and Agricultural Engineering, Kansas 

State University, Manhattan, KS 
ix

Dr. Judith L. Meyer, Distinguished Research Professor Emeritus, Institute of Ecology, 

University of Georgia, Athens, GA 

Dr. James Opaluch, Professor, Department of Environmental and Natural Resource 

Economics, College of the Environment and Life Sciences, University of Rhode Island, 

Kingston, RI 

Dr. Hans Paerl, Professor of Marine and Environmental Sciences, Institute of Marine 

Sciences, University of North Carolina, Chapel Hill, Morehead City, NC 

Dr. Kenneth Reckhow, Professor and Chair, Environmental Science & Policy, Nicholas 

School, Duke University, Durham, NC 

Dr. James Sanders, Director, Skidaway Institute of Oceanography, Savannah, GA 

Dr. Andrew N. Sharpley, Research Soil Scientist, Department of Crop, Soil and 

Environmental Sciences, University of Arkansas, Fayetteville, AR 

Dr. Thomas W. Simpson, Professor and Coordinator, Chesapeake Bay Programs, 

College of Agriculture and Natural Resources, University of Maryland, College Park, 

MD 

Dr. Clifford Snyder, Nitrogen Program Director, International Plant Nutrition Institute, 

Conway, AR 

Dr. Donelson Wright, Chancellor Professor Emeritus, School of Marine Science, 

Virginia Institute of Marine Science, College of William and Mary, Gloucester Point, VA 

SCIENCE ADVISORY BOARD STAFF 

Dr. Holly Stallworth, Designated Federal Officer, EPA Science Advisory Board Staff 

Office, Washington, D.C. 

Dr. Thomas Armitage, Designated Federal Officer, EPA Science Advisory Board Staff 

Office, Washington, D.C. 

Mr. David Wangsness, Designated Federal Officer, Senior Scientist on detail to SAB, 

U.S. Geological Survey, Atlanta, GA 
x

U.S. Environmental Protection Agency 

Science Advisory Board 

CHAIR 

Dr. M. Granger Morgan, Lord Chair Professor in Engineering; Professor and 

Department Head, Department of Engineering and Public Policy, Carnegie Mellon 

University, Pittsburgh, PA 

SAB MEMBERS 

Dr. Gregory Biddinger, Coordinator, Natural Land Management Programs, 

Toxicicology and Environmental Sciences, ExxonMobil Biomedical Sciences, Houston, 

TX 

Dr. Thomas Burke, Professor and Co-Director Risk Sciences and Public Policy Institute, 

Bloomberg School of Public Health, Johns Hopkins University, Baltimore, MD 

Dr. James Bus, Director of External Technology, Toxicology and Environmental 

Research and Consulting, Dow Chemical Company, Midland, MI 

Dr. Deborah Cory-Slechta, J. Lowell Orbison Distinguished Alumni Professor of 

Environmental Medicine, Department of Environmental Medicine, School of Medicine 

and Dentistry, University of Rochester, Rochester, NY 

Dr. Maureen L. Cropper, Professor, Department of Economics, University of 

Maryland, College Park, MD 

Dr. Virginia Dale, Corporate Fellow, Environmental Sciences Division, Oak Ridge 

National Laboratory, Oak Ridge, TN 

Dr. David Dzombak, Professor, Department of Civil and Environmental Engineering, 

Carnegie Mellon University, Pittsburgh, PA 

Dr. Baruch Fischhoff, Howard Heinz University Professor, Department of Social and 

Decision Sciences, Department of Engineering and Public Policy, Carnegie Mellon 

University, Pittsburgh, PA 

Dr. James Galloway, Professor, Department of Environmental Sciences, University of 

Virginia, Charlottesville, VA 

Dr. James K. Hammitt, Professor of Economics and Decision Sciences, Harvard Center 

for Risk Analysis, Harvard University, Boston, MA 

Dr. Rogene Henderson, Scientist Emeritus, Lovelace Respiratory Research Institute, 

Albuquerque, NM 
xi

Dr. James H. Johnson, Professor and Dean, College of Engineering, Architecture & 

Computer Sciences, Howard University, Washington, DC 

Dr. Bernd Kahn, Professor Emeritus and Director, Environmental Resources Center, 

School of Nuclear Engineering and Health Physics, Georgia Institute of Technology, 

Atlanta, GA 

 

Dr. Agnes Kane, Professor and Chair, Department of Pathology and Laboratory 

Medicine, Brown University, Providence, RI 

Dr. Meryl Karol, Professor Emerita, Graduate School of Public Health, University of 

Pittsburgh, Pittsburgh, PA 

Dr. Catherine Kling, Professor, Department of Economics, Iowa State University, 

Ames, IA 

Dr. George Lambert, Associate Professor of Pediatrics, Director, Center for Childhood 

Neurotoxicology, Robert Wood Johnson Medical School-UMDNJ, Belle Mead, NJ 

Dr. Jill Lipoti, Director, Division of Environmental Safety and Health, New Jersey 

Department of Environmental Protection, Trenton, NJ 

 

Dr. Michael J. McFarland, Associate Professor, Department of Civil and 

Environmental Engineering, Utah State University, Logan, UT 

Dr. Judith L. Meyer, Distinguished Research Professor Emeritus, Institute of Ecology, 

University of Georgia, Lopez Island, WA 

Dr. Jana Milford, Associate Professor, Department of Mechanical Engineering, 

University of Colorado, Boulder, CO 

Dr. Rebecca Parkin, Professor and Associate Dean, Environmental and Occupational 

Health, School of Public Health and Health Services, The George Washington University 

Medical Center, Washington, DC 

Mr. David Rejeski, Director, Foresight and Governance Project, Woodrow Wilson 

International Center for Scholars, Washington, DC 

Dr. Stephen M. Roberts, Professor, Department of Physiological Sciences, Director, 

Center for Environmental and Human Toxicology, University of Florida, Gainesville, FL 

Dr. Joan B. Rose, Professor and Homer Nowlin Chair for Water Research, Department 

of Fisheries and Wildlife, Michigan State University 
xii

Dr. James Sanders, Director, Skidaway Institute of Oceanography, University of 

Georgia, Savannah, GA 

Dr. Jerald Schnoor, Allen S. Henry Chair Professor, Department of Civil and 

Environmental Engineering, Co-Director, Center for Global and Regional Environmental 

Research, University of Iowa, Iowa City, IA 

Dr. Kathleen Segerson, Professor, Department of Economics, University of 

Connecticut, Storrs, CT 

Dr. Kristin Shrader-Frechette, O'Neil Professor of Philosophy, Department of 

Biological Sciences and Philosophy Department, University of Notre Dame, Notre Dame, 

IN 

Dr. Philip Singer, Professor, Department of Environmental Sciences and Engineering, 

School of Public Health, University of North Carolina, Chapel Hill, NC 

Dr. Kerry Smith, W.P. Carey Professor of Economics, Dept. of Economics, Carey 

School of Business, Arizona State University, Tempe, AZ 

Dr. Deborah Swackhamer, Interim Director and Professor, Institute on the 

Environment, University of Minnesota, St. Paul, MN 

Dr. Thomas L. Theis, Director, Institute for Environmental Science and Policy, 

University of Illinois at Chicago, Chicago, IL 

Dr. Valerie Thomas, Anderson Interface Associate Professor, School of Industrial and 

Systems Engineering, Georgia Institute of Technology, Atlanta, GA 

Dr. Barton H. (Buzz) Thompson, Jr., Robert E. Paradise Professor of Natural 

Resources Law at the Stanford Law School and Director, Woods Institute for the 

Environment Director, Stanford University, Stanford, CA 

Dr. Robert Twiss, Professor Emeritus, University of California-Berkeley, Ross, CA 

Dr. Lauren Zeise, Chief, Reproductive and Cancer Hazard Assessment Branch, Office 

of Environmental Health Hazard Assessment, California Environmental Protection 

Agency, Oakland, CA 

SCIENCE ADVISORY BOARD STAFF 

Mr. Thomas Miller, EPA Science Advisory Board Staff Office, Washington, DC
xiii

 

Table of Contents 

Table of Figures............................................................................................................. xv 

List of Tables................................................................................................................. xix 

Glossary of Terms......................................................................................................... xxi 

List of Acronyms ......................................................................................................... xxvi 

Conversion Factors and Abbreviations....................................................................... xxx 

Executive Summary....................................................................................................... 1 

1. Introduction........................................................................................................ 10 

1.1. Hypoxia and the Northern Gulf of Mexico – A Brief Overview....................... 10 

1.2. Science and Management Goals for Reducing Hypoxia ................................... 12 

1.3. EPA Science Advisory Board (SAB) Hypoxia Advisory Panel........................ 14 

1.4. The SAB Panel’s Approach............................................................................... 16 

2. Characterization of Hypoxia.............................................................................. 18 

2.1. Processes in the Formation of Hypoxia in the Gulf of Mexico. ........................ 18 

2.1.1. Historical Patterns and Evidence for Hypoxia on the Shelf ..................... 18 

2.1.2. The Physical Context ................................................................................ 21 

2.1.3. Role of N and P in Controlling Primary Production................................. 32 

2.1.4. Other Limiting Factors and the Role of Si................................................ 38 

2.1.5. Sources of Organic Matter to the Hypoxic Zone...................................... 40 

2.1.6. Denitrification, P Burial, and Nutrient Recycling .................................... 46 

2.1.7. Possible Regime Shift in the Gulf of Mexico ........................................... 49 

2.1.8. Single Versus Dual Nutrient Removal Strategies..................................... 52 

2.1.9. Current State of Forecasting ..................................................................... 54 

3. Nutrient Fate, Transport, and Sources ............................................................... 59 

3.1. Temporal Characteristics of Streamflow and Nutrient Flux.............................. 59 

3.1.1. MARB Annual and Seasonal Fluxes ........................................................ 65 

3.1.2. Subbasin Annual and Seasonal Flux......................................................... 73 

3.1.3. Key Findings and Recommendations on Temporal Characteristics......... 83 

3.2. Mass Balance of Nutrients................................................................................. 85 

3.3. Nutrient Transport Processes............................................................................. 98 

3.4. Ability to Route and Predict Nutrient Delivery to the Gulf............................. 107 

4. Scientific Basis for Goals and Management Options ...................................... 120 

4.1. Adaptive Management..................................................................................... 120 

4.2. Setting Targets for Nitrogen and Phosphorus Reduction ................................ 125 

4.3. Protecting Water Quality and Social Welfare in the Basin ............................. 131 

4.4. Cost-Effective Approaches for Non-point Source Control.............................. 143 

4.4.1. Voluntary programs – without economic incentives .............................. 144 

4.4.2. Existing Agricultural Conservation Programs........................................ 145 

4.4.3. Emissions and Water Quality Trading Programs.................................... 147 

4.4.4. Agricultural Subsidies and Conservation Compliance Provisions ......... 148 

4.4.5. Taxes....................................................................................................... 151 

4.4.6. Eco-labeling and Consumer Driven Demand ......................................... 151 

4.4.7. Key Findings and Recommendations on Cost Effective Approaches .... 152 
xiv

4.5. Options for Managing Nutrients, Co-benefits, and Consequences.................. 154 

4.5.1. Agricultural drainage .............................................................................. 154 

4.5.2. Freshwater Wetlands............................................................................... 156 

4.5.3. Conservation Buffers .............................................................................. 161 

4.5.4. Cropping systems.................................................................................... 165 

4.5.5. Animal Production Systems.................................................................... 168 

4.5.6. In-field Nutrient Management ................................................................ 175 

4.5.7. Effective Actions for Other Non-Point Sources ..................................... 195 

4.5.8. Most Effective Actions for Industrial and Municipal Sources ............... 198 

4.5.9. Ethanol and Water Quality in the MARB............................................... 202 

4.5.10. Integrating Conservation Options........................................................... 208 

5. Summary of Findings and Recommendations ................................................. 217 

5.1. Charge Questions on Characterization of Hypoxia ......................................... 217 

5.2. Charge Questions on Nutrient Fate, Transport and Sources............................ 219 

5.3. Charge Questions on Goals and Management Options ................................... 221 

5.4. Conclusion ....................................................................................................... 223 

References.................................................................................................................... 226 

A. Appendix A: Studies on the Effects of Hypoxia on Living Resources ........... A-1 

B. Appendix B: Flow Diagrams and Mass Balance of Nutrients........................ B-8 

C. Appendix C: EPA’s Guidance on Nutrient Criteria...................................... C-15 

D. Appendix D: Calculation of Point Source Inputs of N and P ....................... D-19 

E. Appendix E: Animal Production Systems .....................................................E-21 
xv

Table of Figures 

Figure 1: Map of the frequency of hypoxia in the northern Gulf of Mexico, 1985-2005. Taken from N.N. 

Rabalais, LUMCON, 2006. .......................................................................................................................... 10 

Figure 2: Map showing the extent of the Mississippi-Atchafalaya River basin. ......................................... 11 

Figure 3: Plots of the PEB index (%PEB) in sediment cores from the Louisiana shelf. Higher values of the 

PEB index indicate lower dissolved oxygen contents in bottom waters. Taken from Osterman et al. (2005).

...................................................................................................................................................................... 20 

Figure 4: Change in the relative importance of the Atchafalaya flow to the combined flows from the 

Mississippi and Atchafalaya Rivers over the 20th Century. Reprinted from Bratkovich et al. (1994)........ 24 

Figure 5: Modelled surface salinity showing the freshwater plumes from the Atchafalaya and Mississippi 

Rivers during upwelling favorable winds (top panel) and during downwelling favorable winds 8 days later 

(bottom panel). Adapted from Hetland and DiMarco (2007)....................................................................... 26 

Figure 6: Proposed diversions of Mississippi effluents for coastal protection. From Coastal Protection and 

Restoration Authority (CPRA) of Louisiana, 2007 Integrated Ecosystem Restoration and Hurricane 

Protection: Louisiana’s Comprehensive Master Plan for a Sustainable Coast. CPRA, Office of the Governor 

(La) 117 pp. .................................................................................................................................................. 27 

Figure 7: An illustration depicting different zones (Zones 1-4, numbered above) in the NGOM during the 

period when hypoxia can occur. These zones are controlled by differing physical, chemical, and biological 

processes, are variable in size, and move temporally and spatially. Diagram created by D. Gilbert........... 29 

Figure 8: Response of natural phytoplankton assemblages from coastal NGOM stations to nutrient 

additions, March through September. All experiments, except those done in September, indicate a strong 

response to P additions. Taken from Sylvan et al., 2006............................................................................. 34 

Figure 9: NASA-SeaWiFS image of the Northern Gulf of Mexico recorded in April, 2000. This image 

shows the distributions and relative concentrations of chlorophyll a, an indicator of phytoplankton biomass 

in this region. Note the very high concentrations (orange to red) present in the inshore regions of the 

mouths of the Mississippi and Atchafalaya Rivers....................................................................................... 35 

Figure 10: Estimated extent of agricultural drainage based on the distribution of row crops, largely corn 

and soybean, and poorly drained soils (per D. Jaynes, National Soil Tilth Lab, Ames, IA). ....................... 60 

Figure 11: Land cover based on Landsat data (adapted from Crumpton et al., 2006).................................. 61 

Figure 12: Flow weighted average nitrate concentrations estimated from STORET data selected to exclude 

point source influences (adapted from Crumpton et al., 2006)..................................................................... 61 

Figure 13: Flow-weighted average nitrate and reduced N versus percent cropland (adapted from Crumpton 

et al., 2006)................................................................................................................................................... 62 

Figure 14: MARB nitrate-N fluxes for 1955 through 2005 water years comparing estimates from various 

methods for 1979 to 2005. Based on USGS data from Battaglin (2006) and Aulenbach et al. (2007). ...... 64 

Figure 15: Comparison (percent and absolute basis) of MARB nitrate-N fluxes to LOADEST 5 yr method 

for 1979 through 2005 water years. Based on USGS data from Battaglin (2006) and Aulenbach et al. 

(2007). .......................................................................................................................................................... 65 
xvi

Figure 16: Schematic showing locations of MARB monitoring sites (Aulenbach et al., 2007).................. 66 

Figure 17: Flow and available nitrogen monitoring data for the MARB for 1955 through 2005 water years. 

(LOWESS, Locally Weighted Scatterplot Smooth, curves shown in red). LOWESS describes the 

relationship between Y and X without assuming linearity or normality of residuals, and is a robust 

description of the data pattern (Helsel and Hirsch, 2002). ........................................................................... 67 

Figure 18: Flow, available phosphorus, and available silicate monitoring data for the MARB for 1955 

through 2005 water years. (LOWESS curves shown in red). Based on USGS data from Battaglin (2006) 

and Aulenbach et al. (2007).......................................................................................................................... 68 

Figure 19: Ratio of total N to total P and dissolved silicate to dissolved inorganic N for MARB for the 

1980 through 2005 water years. Based on USGS data from Battaglin (2006) and Aulenbach et al. (2007).

...................................................................................................................................................................... 69 

Figure 20: Flow and nitrogen flux for the MARB during spring (April, May, and June) for the period 1979-

2005. (LOWESS curve shown in red). Based on USGS data from Battaglin (2006) and Aulenbach et al. 

(2007). .......................................................................................................................................................... 70 

Figure 21: Flow, phosphorus, and silicate flux for the MARB during spring (April, May, and June) for the 

period 1979-2006. (LOWESS curve shown in red). Based on USGS data from Battaglin (2006) and 

Aulenbach et al. (2007). ............................................................................................................................... 71 

Figure 22: Sum of April, May and June fluxes as a percent of annual (water year basis) for combined 

Mississippi mainstem and Atchafalaya River. Box plots show median (line in center of box), 25th and 75th 

percentiles (bottom and top of box, respectively), 10th and 90th percentiles (bottom and top error bars, 

respectively) and values < 10th percentile and > 90th percentile (solid circles below and above error bars, 

respectively). Based on USGS data from Battaglin (2006) and Aulenbach et al. (2007)............................ 72 

Figure 23: Ratio of total N to total P and silicate to dissolved inorganic N for the MARB during spring 

(April, May, and June) for the period 1980-2006. Based on USGS data from Battaglin (2006) and 

Aulenbach et al. (2007). ............................................................................................................................... 73 

Figure 24: Location of nine large subbasins comprising the MARB that are used for estimating nutrient 

fluxes (from Aulenbach et al., 2007)............................................................................................................ 74 

Figure 25: Net N inputs and annual nitrate-N fluxes and yields for the Ohio River subbasin. (LOWESS 

curves for riverine nitrate-N shown in red.) Based on USGS data from Battaglin (2006) and Aulenbach et 

al. (2007). ..................................................................................................................................................... 78 

Figure 26: Net N inputs and annual nitrate-N fluxes and yields for the upper Mississippi River subbasin. 

(LOWESS curves for riverine nitrate-N shown in red.) Shown in green is a recalculated net N input for the 

upper Mississippi River basin, increasing soybean N2 fixation from 50 to 70% of above ground N, and a 

soil net N mineralization rate from 0 to 10 kg N/ha/yr. Based on USGS data from Battaglin (2006) and 

Aulenbach et al. (2007). ............................................................................................................................... 79 

Figure 27: Total P and particulate/organic P fluxes for the Ohio River near Grand Chain, Illinois. 

(LOWESS curves shown in black and red). Based on USGS data from Battaglin (2006) and Aulenbach et 

al. (2007). ..................................................................................................................................................... 81 

Figure 28: Spring water flux and nitrate-N flux for the Mississippi River at Grafton and the Ohio River at 

Grand Chain, IL for water years 1975-2005. (LOWESS curves shown in red.) Based on USGS data from 

Battaglin (2006) and Aulenbach et al. (2007)............................................................................................... 82 
xvii

Figure 29: Spring nitrate-N flux (sum of April, May, and June) for the Mississippi River at Grafton plus 

Ohio River at Grand Chain subbasins compared to the combined Mississippi and Atchafalaya River for 

1979 through 2005. Based on USGS data from Battaglin (2006) and Aulenbach et al. (2007). ................. 83 

Figure 30: Area of major crops planted in the MARB from 1941 through 2007. Adapted from McIsaac, 

2006.............................................................................................................................................................. 86 

Figure 31: Nitrogen mass balance components and net N inputs for the MARB, as calculated by McIsaac 

et al. (2002) and updated through 2005 by McIsaac (2006)......................................................................... 88 

Figure 32: Net N inputs for the four major regions of the MARB through 2005. Adapted from McIsaac, 

2006.............................................................................................................................................................. 89 

Figure 33: Nitrogen mass balance components and net N inputs for the upper Mississippi River basin, as 

calculated by McIsaac et al. (2002) and updated through 2005 by McIsaac (2006)..................................... 90 

Figure 34: Phosphorus mass balance components and net P inputs for the MARB. Adapted from McIsaac, 

2006.............................................................................................................................................................. 92 

Figure 35: Net P inputs for the four major subbasins of the MARB through 2005. Adaptive from McIsaac, 

2006.............................................................................................................................................................. 93 

Figure 36: Phosphorus mass balance components and net N inputs for the upper Mississippi River basin. 

Adapted from McIsaac, 2006. ...................................................................................................................... 94 

Figure 37: Total phosphorus point source fluxes as a percent of total flux for the MARB for 2004 by 

hydrologic region. ........................................................................................................................................ 96 

Figure 38: Percentage of nutrient inputs to streams that are removed by instream and reservoir processes as 

predicted by the SPARROW model (Alexander et al., in press). ................................................................. 98

Figure 39: N removed in aquatic ecosystems (as a % of inputs) as a function of ecosystem depth/water 

travel time (modified from David et al., 2006). Values shown are for 23 years in an Illinois reservoir (David 

et al., 2006), French reservoirs (Garnier et al., 1999), Illinois streams (an average from Royer et al., 2004), 

agricultural streams (Opdyke et al., 2006), and rivers (Seitzinger et al., 2002). The curve from Seitzinger et 

al. (2002) is not as steep as the curve that includes information from reservoirs in an agricultural region.100 

Figure 40: A Conceptual Framework for Hypoxia in the Northern Gulf of Mexico................................. 121 

Figure 41: Percent mass nitrate removal in wetlands as a function of hydraulic loading rate. Best fit for 

percent mass loss = 103*(hydraulic loading rate)-0.33. R2 = 0.69. Adapted from Crumpton et al. (2006, in 

press). ......................................................................................................................................................... 158 

Figure 42: Observed NO3 mass removal (blue points) versus predicted NO3 mass removal (blue surface) 

based on the function [mass NO3 removed = 10.3*(HLR) 0.67 * FWA] for which R2 = 0.94. Blue lines are 

isopleths of predicted mass removal at intervals of 250 kg ha/yr. The dashed, red line represents the 

isopleth for mass removal rate of 290 kg ha/yr suggested by Mitsch et al. (2005a). The green plane 

intersecting function surface represents organic N export. Adapted from Crumpton et al. (2006, in press).

.................................................................................................................................................................... 159 

Figure 43: Recoverable manure N, assuming no export of manure from the farm, using 1997 census data. 

Adapted from USDA (2003) with the author’s permission........................................................................ 169 

Figure 44: Recoverable manure P, assuming no export of manure from the farm, using 1997 census data. 

Adapted from USDA (2003) with the author’s permission........................................................................ 170 
xviii

Figure 45: Fertilizer N consumption as anhydrous ammonia in leading corn-producing states for years 

ending June 30............................................................................................................................................ 176 

Figure 46: Changes in the consumption of principal fertilizer N sources used in the six leading cornproducing states (IA, IL, IN, MN, NE, and OH) for years ending June 30. ............................................... 177 

Figure 47: Percentage of N fertilized corn acreage which received some amount of N in the fall............ 178 

Figure 48: USDA ARMS data for the three states with highest fall N application, showing total amount of 

fall applied N for that crop. Also shown are Illinois sales data for the same period. ................................. 179 

Figure 49: Fraction of annual fertilizer N tonnage in Illinois sold in the fall. ........................................... 180

Figure 50: Average corn yields in six leading corn-producing states (IA, IL, IN, MN, NE, and OH), 1990-

2006 (Source:USDA National Agricultural Statistics Service). ................................................................. 183 

Figure 51: Variability in soil test P levels in typical farmer fields in Minnesota (2007 personal 

communication with Dr. Gary Malzer, University of Minnesota).............................................................. 190 

Figure 52: Effect of variable-rate versus uniform rate application of liquid swine manure on changes in soil 

test phosphorus in Iowa fields [2007 personal communication with Dr. Antonio Mallarino, Iowa State 

University and Wittry and Mallarino (2002)]............................................................................................. 191 

Figure 53: Effect of variable rate versus uniform rate application of fertilizer P on soil test P in multiple 

Iowa fields across multiple years (2007 personal communication with A. Mallarino, Iowa State University).

.................................................................................................................................................................... 192 

Figure 54: Nitrogen Cycle Flow Diagram. Taken from Encyclopedia of Earth (2007) at 

http:www.eoearth.org/global_material_cycles. .......................................................................................... B-9 

Figure 55: Phosphorus Cycle Flow Diagram. Taken from Encyclopedia of Earth (2007) at 

http:www.eoearth.org/global_material_cycles. .........................................................................................B-10 

Figure 56: Silicon Cycle Flow Diagram. Taken from Encyclopedia of Earth (2007) at 

http:www.eoearth.org/global_material_cycles. .........................................................................................B-11 

Figure 57: Annual average deposition of NOy across the United States (kg N /hectare-year) based on betatesting runs of the CMAQ model ..............................................................................................................B-14 
xix

List of Tables 

Table 1: A partial summary of papers published following the Integrated Assessment related to sources of 

organic matter to the Gulf of Mexico. .......................................................................................................... 45

Table 2: Site name and corresponding map number for sites discussed in the following section. .............. 75

Table 3: Average annual nutrient fluxes for the five large subbasins in the MARB for the 2001-2005 water 

years. (Percent of total basin flux shown in parentheses.)............................................................................ 75

Table 4: Average annual nutrient fluxes for ten subbasins in the MARB for the 2001-2005 water years.. 

Some subbasin fluxes are calculated as the difference between the upstream and downstream monitoring 

station. (Percent of total basin flux shown in parentheses.).......................................................................... 76

Table 5: Average annual nutrient yields for the five large subbasins in the MARB for water years 2001-

2005.............................................................................................................................................................. 77

Table 6: Average annual nutrient yields for nine subbasins in the MARB for the 2001 - 2005 water years. 

Some subbasin yields are calculated as the difference between the upstream and downstream monitoring 

stations.......................................................................................................................................................... 77

Table 7: Acres of wetlands created, restored or enhanced in major subbasins of the Mississippi River from 

2000-2006 under the Wetland Reserve Program (WRP), Conservation Reserve Program (CRP), 

Conservation Reserve Enhancement Program (CREP), Environmental Quality Incentive Program (EQIP), 

and Conservation Technical Assistance (CTA). (Personal communication, Mike Sullivan, USDA). ....... 104

Table 8. Attributes of models used to estimate sources, transport and/or delivery of nutrients to the Gulf of 

Mexico........................................................................................................................................................ 108

Table 9: Annual and spring (sum of April, May, June) average flow and N and P fluxes for the MARB for 

the 1980 to 1996 reference period compared to the most recent five year period (2001 to 2005). Load 

reductions in mass of N or P also shown.................................................................................................... 127

Table 10: Summary of Study features of Basin wide Integrated Economic-Biophysical Models............. 138

Table 11: Summary of Policies and Findings from Integrated Economic-Biophysical Models................ 139

Table 12: Areas (ha) of conservation buffers installed in the six sub-basins of the MARB for FY 2000 - 

FY2006....................................................................................................................................................... 164

Table 13: Status of implementation of permits under the 2003 CAFO rule for states within the MARB. 

Data provided by EPA Office of Wastewater Management, 2007............................................................. 171

Table 14: Estimates of manure production and N and P loss to water and air from Animal Feeding 

Operations within the Mississippi River basin, on information from the 2002 U.S. Census of Agriculture 

(adapted from Aillery et al., 2005). ............................................................................................................ 173

Table 15: Partial N balance for 4-year rate study by Jaynes et al. (2001). The last two columns added here 

and were not part of original table.............................................................................................................. 186

Table 16: Estimated changes in N losses from cropping changes predicted by FAPRI from 2006-2013. 205

Table 17: Potential total nitrogen (TN) and phosphorus (TP) reduction efficiencies (percent change) in 

surface runoff, subsurface flow, and tile drainage. Estimates are average values for a multiple year basis, 

and some of the numbers in this table are based on a very small amount of field information. ................. 210
xx

Table 18: Anticipated benefits associated with different agricultural management options. .................... 213

Table 19: Anticipated benefits associated with other management options. ............................................. 214

Table 20: Comparison of MART estimated sewage treatment plant annual effluent loads of total N and P 

and values from measurements at each plant for 2004............................................................................. D-20

Table 21: Farming System and Nutrient Budget. .....................................................................................E-22

Table 22: Number of animals and amount of manure produced and N and P excreted within the MARB 

states based on information from the 1997 U.S. Census of Agriculture (data obtained from USDA-ERS, 

http://ers.usda.gov/data/MANURE/). ........................................................................................................E-23
xxi

Glossary of Terms 

Algae: A group of chiefly aquatic plants (e.g., seaweed, pond scum, stonewort, 

phytoplankton) that contain chlorophyll and may passively drift, weakly swim, grow on a 

substrate, or establish root-like anchors (steadfasts) in a water body. 

Anaerobic digestion: Decomposition of biological wastes by micro-organisms, usually 

under wet conditions, in the absence of air (oxygen), to produce a gas comprising mostly 

methane and carbon dioxide. 

Animal feeding operation (AFO): Agricultural enterprises where animals are kept and 

raised in confined situations. AFOs congregate animals, feed, manure and urine, dead 

animals, and production operations on a small land area. Feed is brought to the animals 

rather than the animals grazing or otherwise seeking feed in pastures, fields, or on 

rangeland. Winter feeding of animals on pasture or rangeland is not normally considered 

an AFO. 

Anoxia: The absence of dissolved oxygen. 

Bacterioplankton: The bacterial component of the plankton that drifts in the water 

column. 

Benthic organisms: Organisms living in association with the bottom of aquatic 

environments (e.g., polychaetes, clams, snails). 

Best Management Practices (BMPs): BMPs are effective, practical, structural or 

nonstructural methods that are designed to prevent or reduce the movement of sediment, 

nutrients, pesticides and other chemical contaminants from the land to surface or ground 

water, or which otherwise protect water quality from potential adverse effects of 

agricultural activities. These practices are developed to achieve a cost-effective balance 

between water quality protection and the agricultural production (e.g., crop, forage, 

animal, forest). 

Bioenergy: Useful, renewable energy produced from organic matter - the conversion of 

the complex carbohydrates in organic matter to energy. Organic matter may either be 

used directly as a fuel, processed into liquids and gasses, or be a residual of processing 

and conversion. 

Biogas: A combustible gas derived from decomposing biological waste under anaerobic 

conditions. Biogas normally consists of 50 to 60 percent methane. See also landfill gas. 

Biomass: Any organic matter that is available on a renewable or recurring basis, 

including agricultural crops and trees, wood and wood residues, plants (including aquatic 

plants), grasses, animal residues, municipal residues, and other residue materials. 

Biomass is generally produced in a sustainable manner from water and carbon dioxide by 
xxii

photosynthesis. There are three main categories of biomass - primary, secondary, and 

tertiary. 

Bioreactor: A container in which a biological reaction takes place. As used in this report 

a bioreactor is a container or a trench filled with a biodegradeable carbon source used to 

enhance biological denitrification for removal of nitrate from drainage water. 

Biosolids: Nutrient-rich soil-like materials resulting from the treatment of domestic 

sewage in a treatment facility. During treatment, bacteria and other tiny organisms break 

sewage down into organic matter, sometimes used as fertilizer. 

Cellulosic ethanol: Ethanol that is produced from cellulose material; a long chain of 

simple sugar molecules and the principal chemical constituent of cell walls of plants. 

Chlorophyll: Pigment found in plant cells that are active in harnessing energy during 

photosynthesis. 

Conservation Reserve Program (CRP): CRP provides farm owners or operators with an 

annual per-acre rental payment and half the cost of establishing a permanent land cover, 

in exchange for retiring environmentally sensitive cropland from production for 10- to 

15-years. In 1996, Congress reauthorized CRP for an additional round of contracts, 

limiting enrollment to 36.4 million acres at any time. The 2002 Farm Act increased the 

enrollment limit to 39 million acres. Producers can offer land for competitive bidding 

based on an Environmental Benefits Index (EBI) during periodic signups, or can 

automatically enroll more limited acreages in practices such as riparian buffers, field 

windbreaks, and grass strips on a continuous basis. CRP is funded through the 

Commodity Credit Corporation (CCC). 

Conservation practices (CPs): Any action taken to produce environmental 

improvements, particularly with respect to agricultural non-point source emissions. The 

term is used broadly to refer to structural practices, such as buffers, as well as 

nonstructural preactices, such as in-field nutrient management planning and application. 

Conservation Practice standards have been developed by NRCS and are available at: 

http://www.nrcs.usda.gov/Technical/Standards/nhcp.html. 

Corn stover: Corn stocks that remain after the corn is harvested. Such stocks are low in 

water content and very bulky. 

Cyanobacteria: A phylum (or “division”) of bacteria that obtain their energy through 

photosynthesis. They are often referred to as blue-green algae, although they are in fact 

prokaryotes, not algae. The description is primarily used to reflect their appearance and 

ecological role rather than their evolutionary lineage. The name “cyanobacteria” comes 

from the color of the bacteria, cyan. 

Demersal organisms: Organisms that are, at times, associated with the bottom of aquatic 

environments, but capable of moving away from it (e.g., blue crabs, shrimp, red drum). 
xxiii

Denitrification: Nitrogen transformations in water and soil that make nitrogen effectively 

unavailable for plant uptake, usually returning it to the atmosphere as nitrogen gas. 

Diatom: A major phytoplankton group characterized by cells enclosed in silicon 

frustules, or shells. 

Dinoflagellates: Mostly single-celled photosynthetic algae that bear flagella (long cell 

extensions that function in swimming) and live in fresh or marine waters. 

Edge-of-field nitrogen loss: A term that refers to the nitrogen that is lost or exported 

from fields in agricultural production. 

Effluent: The liquid or gas discharged from a process or chemical reactor, usually 

containing residues from that process. 

Emissions: Waste substances released into the air or water. See also Effluent. 

Eutrophic: Waters, soils, or habitats that are high in nutrients; in aquatic systems, 

associated with wide swings in dissolved oxygen concentrations and frequent algal 

blooms. 

Eutrophication: An increase in the rate of supply of organic matter to an ecosystem. 

Greenhouse gases: Gases that trap the heat of the sun in the Earth's atmosphere, 

producing the greenhouse effect. The two major greenhouse gases are water vapor and 

carbon dioxide. Other greenhouse gases include methane, ozone, chlorofluorocarbons, 

and nitrous oxide. 

Hydrogen sulfide: A chemical, toxic to oxygen-dependent organisms, that diffuses into 

the water as the oxygen levels above the seabed sediments become zero. 

Hypoxia: Very low dissolved oxygen concentrations, generally less than 2 milligrams 

per liter. 

Lignocellulose: A combination of lignin and cellulose that strengthens woody plant cells. 

Nitrate: An inorganic form of nitrogen; chemically NO3. 

Nitrogen fixation: The transformation of atmospheric nitrogen into nitrogen compounds 

that can be used by growing plants. 

Non-point source: A diffuse source of chemical and/or nutrient inputs not attributable to 

any single discharge (e.g., agricultural runoff, urban runoff, atmospheric deposition). 
xxiv

Nutrients: Inorganic chemicals (particularly nitrogen, phosphorus, and silicon) required 

for the growth of plants, including crops and phytoplankton. 

Phytoplankton: Plant life (e.g., algae), usually containing chlorophyll, that passively 

drifts in a water body. 

Plankton: Organisms living suspended in the water column, incapable of moving against 

currents. 

Point source: Readily identifiable inputs where treated wastes are discharged from 

municipal, industrial, and agricultural facilities to the receiving waters through a pipe or 

drain. 

Pre-sidedress-nitrate test (PSNT): A soil nitrate-N test determined in surface soil 

samples (usually 0 to 30 cm or 0 to 12 in deep), collected between corn rows when the 

corn is about 15 cm (6 in) tall. Adjustments in the rate of side-dressed N can be made if 

the soil test indicates elevated nitrate-N levels, based upon calibrations that vary among 

growing regions. When successfully calibrated, the test results can be used as an index of 

the amount of N that may be released during the course of the growing season by organic 

sources such as soil organic matter, manure, and crop residues. 

Productivity: The conversion of light energy and carbon dioxide into living organic 

material. 

Pycnocline: The region of the water column characterized by the strongest vertical 

gradient in density, attributable to temperature, salinity, or both. 

Recoverable manure: The portion of manure as excreted that could be collected from 

buildings and lots where livestock are held, and thus would be available for land 

application. 

Recoverable manure nutrients: The amounts of nitrogen and phosphorus in manure that 

would be expected to be available for land application. They are estimated by adjusting 

the quantity of recoverable manure for nutrient loss during collection, transfer, storage, 

and treatment; but are not adjusted for losses of nutrients at the time of land application. 

Respiration: The consumption of oxygen during energy utilization by cells and 

organisms. 

Riparian floodplain: Area adjacent to a river or other body of water subject to frequent 

flooding. 

Soil tilth: The physical condition of the soil as related to its ease of tillage, fitness as a 

seedbed, and impedance to seedling emergence and root penetration. A soil with good 

“tilth” has large pore spaces for adequate air infiltration and water movement, and holds a 

reasonable supply of water and nutrients. Soil tilth is a factor of soil texture, soil 
xxv

structure, and the interplay with organic content and the living organisms that help make 

up the soil ecosystem. 

Stratification: A multilayered water column, delineated by pycnoclines. 

Sustainable: An ecosystem condition in which biodiversity, renewability, and resource 

productivity are maintained over time. 

Urease and nitrification inhibitors: Urease is a ubiquitous soil microbial enzyme that 

facilitates the hydrolysis of urine and urea to form ammonia. In the soil, ammonia 

readily hydrolyzes to ammonium. Soil ammonium also is formed by the mineralization 

of soil organic matter and manures. Ammonium is then oxidized or “nitrified” first to 

nitrite (NO2) and then to nitrate (NO3), which is highly soluble and subject to movement 

in the soil with the moisture front, or leaching under certain conditions. Under anaerobic 

conditions, NO3 can be “denitrified” to the gases nitrous oxide (N2O) and nitrogen (N2), 

and released to the atmosphere. Urease inhibitors are chemicals applied to fertilizers or 

manures to reduce urease activity. Under certain environmental conditions urease 

inhibitors can temporarily inhibit or reduce ammonia loss (volatilization) to the 

atmosphere from urea-containing fertilizers or manures. Nitrification inhibitors are 

chemicals which can temporaritly inhibit or reduce nitrification of anhydrous ammonia, 

ammonium-containing or urea-containing fertilizers applied to the soil; which may 

indirectly help to reduce denitrification losses of N. Under certain environmental 

conditions, urease and nitrification inhibitors help improve soil retention and crop 

recovery of applied N, which may reduce potential environmental N losses. 

Voluntary programs: Voluntary conservation programs that have no significant financial 

incentive (positive or negative) to encourage the adoption of conservation practices. 

Watershed: The drainage basin contributing water, organic matter, dissolved nutrients, 

and sediments to a stream or lake. 

Zooplankton: Animal life that drifts or weakly swims in a water body, often feeding on 

phytoplankton. 
xxvi

List of Acronyms 

ADCPs – Acoustic Doppler Current Profilers 

AFO – Animal Feeding Operation 

AMLE – Adjusted Maximum Likelihood Estimate 

ANNAMOX – Anaerobic Ammonia Oxidation 

A/P ratio – Agglutinated to Porcelaneous ratio (based on the relative abundance of three 

low-oxygen tolerant species of benthic foraminifers; Pseudononin altlanticum, 

Epistominella vitrea, and Buliminella morgani) 

ARS – Agricultural Research Service (USDA) 

AUs – Animal Units 

BBL – Benthic Boundary Layer 

BMPs – Best Management Practices 

BNR – Biological Nutrient Removal 

BOD – Biochemical Oxygen Demand 

Bu/A – Bushels per acre 

C – Carbon 

CAFO – Concentrated Animal Feeding Operation 

CASTnet – Clean Air Status and Trends Network 

CC or Ccc – Continuous Corn 

CCC – Commodity Credit Corporation 

CCOA – Corn-Corn-Oat-Alfalfa (crop rotation) 

CDOM – Colored Dissolved Organic Matter 

CEAP - Conservation Effectiveness Assessment Program 

CENR – Committee on Environment and Natural Resources 

Cm – Corn-meadow (crop rotation) 

CMAQ – Community Multiscale Air Quality model 

COAA – Corn-Oat-Alfalfa-Alfalfa (crop rotation) 

CO2 – Carbon Dioxide 

cph – cycles per hour 

CPRA – Coastal Protection and Restoration Authority 

CREP – Conservation Reserve Enhancement Program 

CRN – Controlled – and slow Release N fertilizers 

CRP - Conservation Reserve Program 

CRPA – Coastal Protection and Restoration Authority 

CS or CSb – Corn Soybean rotation 

CSP – Conservation Security Program 

CTA – Conservation Technical Assistance 

CTDs – Conductivity, Temperature, and Depth instrumentation 

CVs – Coefficients of Variations 

DDGs – Dried Distillers Grain 

DIN:DIP – Dissolved Inorganic Nitrogen:Dissolved Inorganic Phosphorus 

DO – Dissolved Oxygen 

DOC – Dissolved Organic Carbon 

DOE – Department of Energy 
xxvii

DOM – Dissolved Organic Matter 

DON – Dissolved Organic Nitrogen 

DRP – Dissolved Reactive Phosphorus 

EBI – Environmental Benefits Index 

ECa – Electrical Conductivity 

ENR – Enhanced Nutrient Removal 

EPC0 – Equilibrium P Concentration 

EPIC – Environment Productivity Impact Calculator model 

EQIP -- Environmental Quality Incentives Program 

ERS – Economic Research Service (USDA) 

Fe+2 – Ferrous Iron 

FR – Federal Register 

FWA – Flow Weighted Average 

GAO – General Accounting Office 

GCOOS – Gulf of Mexico Coastal Ocean Observing System 

GCTM – Global Chemistry Transport Model 

GHG – Green House Gases 

GIS – Geographic Information System 

GLWQA – Great Lakes Water Quality Agreement 

GOM -Gulf of Mexico 

GPS – Global Positioning System 

GWW – Grass Waterways 

HAB – Harmful Algal Bloom 

HAP – Hypoxia Advisory Panel or SAB Panel 

HEL – Highly Erodable Land 

HLR – Hydraulic Loading Rate 

HRUs – Hydraulic Response Units 

HUC – Hydrologic Unit Code 

HYDRA – Hydrological Routing Algorithm 

IATP – Institute of Agricultural and Trade Policy 

IBIS – Integrated Biosphere Simulator model 

IJC – International Joint Commission 

IPCC – Intergovernmental Panel on Climate Change 

ISNT – Illinois Soil Nitrogen Test 

LOADEST – Load Estimator model 

LOWESS – Locally Weighted Scatterplot Smooth curves 

LSNT – Late Spring Nitrate Test 

LUMCON – Louisiana Universities Marine Consortium 

M – Million 

MGD – Million gallons per day 

MARB – Mississippi-Atchafalaya River basin 

MART -- Management Action Reassessment Team 

Mn+2 – Manganese (oxidation state common in aquatic-biological systems) 

MRB – Mississippi River basin 

MR/GMWNTF – Mississippi River/Gulf of Mexico Watershed Nutrient Task Force 

MSEA – Management System Evaluation Area 
xxviii

N -- Nitrogen 

N2 – Nitrogen gas (colorless, odorless, and tasteless gas that makes up 78.09% of air) 

N2O – Nitrous Oxide 

NADP – National Air Deposition Program 

NANI -- Net Anthropogenic Nitrogen Inputs 

NAS – National Academy of Sciences 

NASA – National Aeronautics and Space Administration 

NASA-SeaWiFS – NASA Sea-viewing Wide Field-of-view Sensor (project providing 

qualitative data on global ocean bio-optical properties) 

NASQAN – National Stream Quality Accounting Network (USGS water-quality 

monitoring program) 

NECOP – Nutrient Enhanced Coastal Ocean Productivity 

NGOM – Northern Gulf of Mexico 

NH3 -- Ammonia 

NH4

+

 -- Ammonium 

NHx – The total atmospheric concentration of ammonia (NH3) and ammonium (NH4

+

)

NOAA – National Oceanic and Atmospheric Administration 

NO2 – Nitrite Nitrogen (NO2

-

) if in water and Nitrogen Dioxide (NO2) if in air 

NO3 – Nitrate nitrogen 

NOx – Mono-nitrogen oxides, or the total concentration of nitric oxide (NO) plus 

nitrogen dioxide (NO2) 

NOy – Reactive odd nitrogen or the sum of NOx plus compounds produced from the 

oxidation of NOX, which includes nitric acid, peroxyacetyl nitrate, and other compounds 

NPDES – National Pollutant Discharge Elimination System 

NPSs – Non-Point Sources 

NRC – National Research Council 

NRCS – Natural Resource Conservation Service 

NRI – National Resources Inventory 

NSTC – National Science and Technology Council 

O2 – Diatomic Oxygen (makes up 20.95% of air) 

OM – Organic Matter 

P – Phosphorus 

PEB index – An index based on the relative abundance of three low-oxygen tolerant 

species of benthic foraminifers; Pseudononin altlanticum, Epistominella vitrea, and 

Buliminella morgani

POC – Particulate Organic Carbon 

ppmv – Parts per million by volume 

ppt – Parts per thousand 

PS – Point Source 

PSNT – Pre-Sidedress Nitrate Test 

RivR-N -- A regression model that predicts the proportion of N removed from streams 

and reservoirs as an inverse function of the water displacement time of the water body 

(ratio of water body depth to water time of travel)

SAB – Science Advisory Board 

SCOPE – Science Committee on Problems of the Environment 

SD – Standard Deviation 
xxix

Si – Silicon 

SOC – Soil Organic Carbon 

SOM – Soil Organic Matter 

SON – Soil Organic Nitrogen 

SPARROW - Spatially Referenced Regression on Watershed attributes model 

SRP or DRP or ortho P – Soluble Reactive Phosphorus, Dissolved Reactive Phosphorus, 

Orthophosphate 

STATSGO – State Soil Geographic database 

STORET – STOrage and RETrieval data system (EPA’s largest computerized 

environmental data system) 

STPs – Sewage Treatment Plants 

SWAT - Soil and Water Assessment Tool model 

THMB – Terrestrial Hydrology Model with Biogeochemistry 

TKN – Total Kjeldahl Nitrogen 

TM3 – Tracer Model version 3 (a global atmospheric chemistry/transport model) 

TN – Total Nitrogen 

TP – Total Phosphorus 

TPCs – Typical Pollutant Concentrations 

TSS – Total Suspended Solids 

UAN – Urea Ammonium Nitrate 

UMRB – Upper Mississipppi River basin 

UMRSHNC – Upper Mississippi River Sub-basin Hypoxia Nutrient Committee 

USMP – U.S. Agriculture Sector Mathematical Programming model

USACE – United States Army Corps of Engineers 

USDA – United States Department of Agriculture 

USEPA or EPA – United States Environmental Protection Agency 

USGS – United States Geological Survey 

WRP – Wetlands Reserve Program 
xxx

Conversion Factors and Abbreviations

MULTIPLY BY TO OBTAIN 

centimeter (cm) 0.3937 inch (in) 

millimeter (mm) 0.0394 inch (in) 

meter (m) 3.281 foot (ft) 

kilometer (km) 0.6214 mile (mi) 

square kilometer (km2

) 0.3861 square mile (mi2) 

hectare (ha) 2.471 acre (ac) 

hectare (ha) 0.01 square kilometer (km2

) 

liter (L) 1.057 quart (qt) 

liter (L) 0.0284 bushel (bu) US, dry 

gram (g) 0.0353 ounce (oz) 

gram per cubic meter (g/m3

) 0.00169 pound per cubic yard (lb/yd3) 

kilogram (kg) 2.205 pound (lb), avoirdupois 

metric tonne (tonne) 2,205.0 pound (lb), avoirdupois 

metric tonne (tonne) 1.1023 U.S. short ton (ton) 

cubic meter per second (m3/s) 35.31 cubic foot per second (cfs) 

kilogram per hectare (kg/ha) 0.893 pound per acre (lb/ac) 

CONCENTRATION UNIT APPROXIMATELY EQUALS 

milligram per liter (mg/L) part per million (ppm) 

The following equation was used to compute flux of chemicals: 

concentration (mg/L) x flow (m3

/s) x 8.64 x 10-2 = metric tonne per day (tonne/d)
1

Executive Summary 

Since 1985, scientists have been documenting a hypoxic zone in the Gulf of 

Mexico each year. The hypoxic zone, an area of low dissolved oxygen that cannot 

support marine life, generally manifests itself in the spring. Since marine species either 

die or flee the hypoxic zone, the spread of hypoxia reduces the available habitat for 

marine species, which are important for the ecosystem as well as commercial and 

recreational fishing in the Gulf. Since 2001, the hypoxic zone has averaged 16,500 km2 

during its peak summer months1 and ranged from a low of 8,500 km2 to a high of 22,000 

km2

. To address the hypoxia problem, the Mississippi River/Gulf of Mexico Watershed 

Nutrient Task Force (or Task Force) was formed to bring together representatives from 

federal agencies, states and tribes to consider options for responding to hypoxia. The 

Task Force asked the White House Office of Science and Technology Policy to conduct a 

scientific assessment of the causes and consequences of Gulf hypoxia through its 

Committee on Environment and Natural Resources (CENR). In 2000 the CENR 

completed An Integrated Assessment: Hypoxia in the Northern Gulf of Mexico 

(Integrated Assessment), which formed the scientific basis for the Task Force’s Action 

Plan for Reducing, Mitigating and Controlling Hypoxia in the Northern Gulf of Mexico

(Action Plan, 2001). In its Action Plan, the Task Force pledged to implement ten 

management actions and to assess progress every five years. This reassessment would 

address the nutrient load reductions achieved, the responses of the hypoxic zone and 

associated water quality and habitat conditions, and economic and social effects. The 

Task Force began its reassessment in 2005. 

 In 2006 as part of the reassessment, EPA’s Office of Water, on behalf of the Task 

Force, requested that the Environmental Protection Agency (EPA) Science Advisory 

Board (SAB) convene an independent panel to evaluate the state of the science regarding 

hypoxia in the Northern Gulf of Mexico and potential nutrient mitigation and control 

options in the Mississippi-Atchafalaya River basin (MARB). The Task Force was 

particularly interested in scientific advances since the Integrated Assessment and issued 

charge questions in three areas: characterization of hypoxia; nutrient fate, transport and 

sources; and the scientific basis for goals and management options. The SAB Hypoxia 

Advisory Panel (SAB Panel) began its deliberations in September of 2006 and completed 

its report in August of 2007 while operating under the “sunshine” requirements of the 

Federal Advisory Committee Act, which include providing public access to advisory 

meetings and opportunities for public comment. This Executive Summary summarizes

the SAB Panel’s major findings and recommendations. 

 

1

 The areal extent of the full hypoxic region has not been mapped with sufficient frequency to completely 

understand its temporal variability. The limited number of observations that have been taken more than 

once per year suggest that the hypoxic region reaches its maximum extent in late summer. There are 

physical and biological reasons to expect such a pattern of temporal variation but available data provide a 

conservative estimate of the maximum extent of hypoxia. The actual areal extent may be larger than 

estimated. 
2

Findings 

 Since publication of the Integrated Assessment, scientific understanding of the 

causes of hypoxia has grown while actions to control hypoxia have lagged. Recent 

science has affirmed the basic conclusion that contemporary changes in the hypoxic area 

in the northern Gulf of Mexico (NGOM) are primarily related to nutrient fluxes from the 

MARB. Moreover, new research provides early warnings about the deleterious longterm effects of hypoxia on living resources in the Gulf. 

 The SAB Panel was asked to comment on the Action Plan’s goal to reduce the 

hypoxic zone to a five-year running average of 5,000 km2

 by 2015. The 5,000 km2 target 

remains a reasonable endpoint for continued use in an adaptive management context; 

however, it may no longer be possible to achieve this goal by 2015 for two reasons. First, 

there is limited current movement to implement policies, programs and strategies that 

reduce nutrients. Second, there are time lags between reductions in nutrients and the 

response of the ecological system. In August of 2007, the hypoxic zone was measured to 

be 20,500 km2 (LUMCON, 2007), the third largest hypoxic zone since measurements 

began in 1985. Accordingly, it is even more important to proceed in a directionally 

correct fashion to manage factors affecting hypoxia than to wait for greater precision in 

setting the goal for the size of the zone. Much can be learned by implementing 

management plans, documenting practices, and measuring their effects with appropriate 

monitoring programs. 

 To reduce the size of the hypoxic zone and improve water quality in the MARB, 

the SAB Panel recommends a dual nutrient strategy targeting at least a 45% reduction in 

riverine total nitrogen flux (to approximately 870,000 metric tonne/yr or 960,000 ton/yr) 

and at least a 45% reduction in riverine total phosphorus flux (to approximately 75,000 

metric tonne/yr or 83,000 ton/yr). Both of these reductions refer to changes measured 

against average flux over the 1980 - 1996 time period. For both nutrients, incremental 

annual reductions will be needed to achieve the 45% reduction goals over the long run. 

For nitrogen, the greatest emphasis should be placed on reducing spring flux, the time 

period most correlated with the size of the hypoxic zone. While the state of predictive 

and process models of NGOM hypoxia has continued to develop since 2000, models 

similar to those in place at that time are still the best tools for producing dose response

estimates for nitrogen (N) reductions, with most recent model runs showing a 45 – 55% 

required reduction for N in order to reduce the size of the hypoxic zone. A number of 

studies have suggested that climate change will create conditions for which larger 

nutrient reductions, e.g., 50 – 60% for nitrogen, would be required to reduce the size of 

the hypoxic zone. 

 New information has emerged that more precisely demonstrates the role of 

phosphorus (P) in determining the size of the hypoxic zone. Contrary to conventional 

wisdom that N typically limits phytoplankton production in near-coastal waters, the 

NGOM exhibits an unusual phenomenon whereby P is an important limiting constituent 
3

during the spring and summer in the lower salinity, near-shore regions. Phosphorus 

limitation is now occurring because over the past 50 years excessive N loadings have 

dramatically altered nitrogen to phosphorus ratios. Taken together, N and P both 

contribute to excess phytoplankton production and the hypoxia associated with such 

production, and they will need to be reduced concurrently to make progress in reducing 

the size of the hypoxic zone. The SAB Panel’s best professional judgment is that 

phosphorus reductions will need to be comparable (in percentage terms) to nitrogen 

reductions to reduce the size of the hypoxic zone. 

 Scientific advances have improved our understanding of the physical factors that 

contribute to hypoxia. One physical factor that has changed substantially over the past 

century is river hydrology. The hydrologic regime of the Mississippi and Atchafalaya 

Rivers and the timing of freshwater inputs to the continental shelf are critical to mixing 

and hypoxia development. The most important hydrological change over the past century 

has been the diversion of a large amount of freshwater from the Mississippi River 

through the Atchafalaya River to the Atchafalaya Bay, and maintenance of this diversion 

by the U. S. Army Corps of Engineers. The major injection of freshwater into 

Atchafalaya Bay, some 200 kilometers to the west of the Mississippi River Delta, has 

profoundly modified the spatial distribution of freshwater inputs, nutrient loadings and 

stratification on the Louisiana-Texas continental shelf. 

Methods used by the U.S. Geological Survey (USGS) to calculate nutrient fluxes 

in the MARB have changed since the Integrated Assessment. The latest USGS estimates 

show that total N flux averaged 1.24 million metric tonne/yr (1.37 million ton/yr) from 

2001 – 2005 (65% of the flux is nitrate), and the total P flux averaged 154,000 metric 

tonne/yr (170,000 ton/yr). This change represents a 21% decline in total N flux and a 

12% increase in total P flux when compared to the averages from the 1980 – 1996 time 

period. The spring (April – June) flux of nutrients appears to be an important 

determinant of hypoxia, for that is when the river is disproportionately enriched with both 

N (especially nitrate) and P. Spring total N flux has declined since the 1980s; whereas 

total P flux shows a 9.5% increase (when average total P flux for 2001-2005 is compared 

to the 1980 – 1996 average). USGS data also show that during the last 5 years, the upper 

Mississippi and Ohio-Tennessee River subbasins contributed about 82% of nitrate-N 

flux, 69% of the TKN flux, and 58% of total P flux, although these sub-basins represent 

only 31% of the entire MARB drainage area. 

The SAB Panel’s estimates of point source discharge show that point sources 

represented 22% of total annual average N flux and 34% of total annual average P flux 

discharged to the NGOM during the last five years. New methods also have been used to 

calculate nutrient mass balances (net anthropogenic N inputs, NANI). NANI for the 

MARB has declined in the past decade because of increased crop yields, reduced or 

redistributed livestock populations, and little change in N fertilizer inputs. From 1999-

2005, NANI calculations show 54% of non-point N inputs in the MARB were from 

fertilizer, 37% from nitrogen fixation, and 9% from atmospheric deposition. 
4

 The SAB Panel finds that the Gulf of Mexico ecosystem appears to have gone 

through a regime shift with hypoxia such that today the system is more sensitive to inputs 

of nutrients than in the past, with nutrient inputs inducing a larger response in hypoxia as 

shown for other coastal marine ecosystems such as the Chesapeake Bay and Danish 

coastal waters. Changes in benthic and fish communities with the change in frequency 

of hypoxia are cause for concern. The recovery of hypoxic ecosystems may occur only 

after long time periods or with further reductions in nutrient inputs. If actions to control 

hypoxia are not taken, further ecosystem impacts could occur within the Gulf, as has been 

observed in other ecosystems. 

Certain aspects of the nation’s current agricultural and energy policies are at odds 

with the goals of hypoxia reduction and improving water quality. Since the Integrated 

Assessment, an emerging national strategy on renewable fuels has granted economic 

incentives to corn-based ethanol production. The projected increase in corn production 

from this strategy has profound implications for water quality in the MARB, as well as 

hypoxia in the NGOM. Recent energy policies, combined with pre-existing crop 

subsidies, tax policies, global market conditions and trade barriers all provide economic 

incentives for conversion of retired and other cropland to corn production for use in 

ethanol production. Such conversions are projected to lead to corn production on an 

additional 6.5 million ha (16 million ac) in coming years with the majority of this 

increase occurring in the MARB. Without some change to the current structure of 

economic incentives favoring corn-based ethanol, N loadings to the MARB from 

increased corn production could increase dramatically in coming years, rather than 

decreasing, as needed for the NGOM. 

Recommendations for Monitoring and Research 

 Most of the research and monitoring needs identified in the Integrated Assessment

have not been met, and fewer rivers and streams are monitored today than in 2000. The 

majority of monitoring recommendations in the Integrated Assessment remain relevant 

and should be heeded. The SAB Panel affirms and reiterates the CENR’s call to improve 

and expand monitoring of the temporal and spatial extent of hypoxia and the processes 

controlling its formation; the flux of nutrients, carbon, and other constituents from nonpoint sources throughout the MARB and to the NGOM; and measured (rather than 

estimated) nitrogen and phosphorus fluxes from municipal and industrial point sources. 

 The SAB Panel affirms the need for research in the following areas identified in 

the Integrated Assessment: ecological effects of hypoxia; watershed nutrient dynamics; 

effects of different agricultural practices on nutrient losses from land, particularly at the 

small watershed scale; nutrient cycling and carbon dynamics; long-term changes in 

hydrology and climate; and economic and social impacts of hypoxia. 

A suite of models is needed to simulate the processes and linkages that regulate 

the onset, duration and extent of hypoxia. Emerging coastal ocean observation and 

prediction systems should be encouraged to monitor dissolved oxygen and other physical 

and biogeochemical parameters needed to continue improving hypoxia models. 
5

To advance the science characterizing hypoxia and its causes, the SAB Panel 

finds that research is also needed to: 

• collect and analyze additional sediment core data needed to develop a 

better understanding of spatial and temporal trends in hypoxia; 

• investigate freshwater plume dispersal, vertical mixing processes and 

stratification over the Louisiana-Texas continental shelf and Mississippi 

Sound, and use three-dimensional hydrodynamic models to study the 

consequences of past and future flow diversions to NGOM distributaries; 

• advance the understanding of biogeochemical and transport processes 

affecting the load of biologically available nutrients and organic matter to 

the Gulf of Mexico, and develop a suite of models that integrate physics 

and biogeochemistry; 

• elucidate the role of P relative to N in regulating phytoplankton production 

in various zones and seasons, and investigate the linkages between inshore 

primary production, offshore production, and the fate of carbon produced 

in each zone; 

• improve models that characterize the onset, volume, extent, and duration 

of the hypoxic zone, and develop modeling capability to capture the 

importance of P, N, and P-N interactions in hypoxia formation; 

 To advancing the science on sources, fate and transport of nutrients, the SAB 

Panel recommends research to: 

• develop models to simulate fluvial processes and estimate N and P transfer 

to stream channels under different management scenarios; 

• improve the understanding of temporal and seasonal nutrient fluxes and 

develop nutrient, sediment, and organic matter budgets within the MARB; 

 To enhance the scientific basis for implementation of management options, the 

SAB Panel finds that research is needed to: 

• examine the efficacy of dual nutrient control practices; 

• determine the extent, pattern, and intensity of agricultural drainage as well 

as opportunities to reduce nutrient discharge by improving drainage 

management; 

• integrate monitoring, modeling, experimental results, and ongoing 

management into an improved conceptual understanding of how the forces 
6

at key management scales influence the formation of the hypoxia zone; 

and 

• develop integrated economic and watershed models to fully assess costs 

and benefits, including co-benefits, of various management options. 

 Developments in the biofuels industry have created new questions for researchers 

to address. More research is needed on biofuel life cycles in order to identify system 

efficiency with respect to environmental effects, economics, and resource availability of 

biofuel alternatives. That is, research needs to evaluate the environmental effects of 

different biofuel production processes on soil, water quality and climate under realistic 

strategies of deploying production facilities and moving the biofuels to the market. 

Current incentives favor corn-based ethanol production, although research has thus far 

shown fewer environmental consequences with other feedstocks, e.g., cellulosic 

feedstocks such as switchgrass. Yet the technology for conversion of cellulosic 

feedstocks to biofuel is not yet commercially viable. Policies of all kinds (taxes, 

subsidies, trade) could be used to support research and technological developments for 

those biofuels that balance high energy yields with the lowest environmental impacts. 

Recommendations for Adaptive Management 

 Adaptive management provides a framework for ongoing management in the face 

of uncertainty. It requires that conceptual models be developed to guide management and 

that management actions be treated like well-monitored experiments that answer 

questions for improving decisions with each successive cycle of learning. The most 

urgent need is to decrease nutrient discharge. In fact, nutrients should be decreased as 

soon as possible before the system requires even larger nutrient reductions to reduce the 

area of hypoxia. Already many taxa are lost during the peak of hypoxia, and there has 

been a shift in the relative abundance of fish species. Increases in certain pelagic species 

can disrupt food web structure, and the new system may respond in a quite different way 

to changes in nutrient level. The SAB Panel thus agrees with the CENR’s emphasis on 

decreasing nutrient discharge in the context of adaptive management. 

 These adaptive management actions must be interpreted in view of both field 

measures and models of their effects. Conceptual models are needed for nutrient 

management at several spatial resolutions from small catchments, to large watersheds, to 

the entire MARB in order to guide research and ongoing adaptive management at each of 

the relevant scales. To the greatest extent possible, feedbacks should be incorporated into 

the models so that management is accompanied by learning about the full systems of 

linkages between human activities and hypoxia as well as the full range of co-benefits of 

N and P reductions. 
7

Management Options 

 Large N and P reductions, on the order of 45% or more, are needed to reduce the 

size of the hypoxic zone. To do this, the SAB Panel found the most significant 

opportunities for N and P reductions occur in five areas: 

 promotion, via research and economic incentives, of environmentally 

sustainable approaches to biofuel production and associated cropping systems 

(e.g., perennials). 

 improved management of nutrients by emphasizing infield nutrient 

management efficiency and effectiveness to reduce losses; 

 construction and restoration of wetlands, as well as criteria for targeting those 

wetlands that may have a higher priority for reducing nutrient losses; 

 introduction of tighter N and P limits on municipal and private industrial point 

sources; and 

 improved targeting of conservation buffers, including riparian buffers, filter 

strips and grassed waterways, to control surface-borne nutrients. 

Importantly, not all approaches will be cost-effective in all locations; the optimal 

combination and location of these practices will vary across and within watersheds. 

 In terms of cropping systems, research comparing nutrient discharge between 

alternative cropping systems (including row crops and non-row crops such as perennials) 

and a corn-soybean rotation shows that significant nutrient loss reductions could be 

achieved by converting current corn-soybean rotations to alternative crops or alternative 

rotations. Moreover, since corn crops require more nitrogen input, cellulosic sources 

(e.g., perennial grasses, fast-growing woody species, etc.) could, by comparison, provide 

alternative energy while protecting water quality. However, the technology for 

converting cellulosic sources to biofuel is not yet commercially viable. Significant 

reductions in nutrient runoff could also be achieved if nutrients are managed more 

efficiently on farms, for example by moving to spring fertilization rather than fall. More 

wetlands are needed, especially in those areas that promise the greatest N and P 

reductions. Since the greatest N and P runoff is coming from upper Mississippi and 

Ohio-Tennessee River subbasins, where the highest proportion of tile drainage occurs, 

measures to improve drainage water management are urgently needed. In fact, improved 

targeting of almost all agricultural conservation practices in the region [e.g., conservation 

buffers, wetlands, land set aside in the Conservation Reserve Program (CRP), drainage 

water management, etc.] could achieve greater local water quality benefits and 

simultaneously contribute to hypoxia reduction. Nearly all of these opportunities were 

recognized in the Integrated Assessment. 
8

 The CENR did not emphasize tighter limits on municipal point sources; however 

new calculations from the SAB Panel indicate that 22% of annual average total N flux 

and 34% of annual average total P flux to the Gulf comes from permitted point-source 

dischargers. The SAB Panel’s calculations further demonstrate that tighter limits on N 

and P in effluent (3 mg N/L and 0.3 mg P/L) from sewage treatment plants could realize 

an estimated 11% reduction in annual average total N flux and a 21% reduction in total 

annual average P flux to the Gulf. Although the exact N and P limit could be debated, 

clearly there are regulatory opportunities to significantly reduce N and P fluxes to the 

Gulf. The cost associated with such regulations could be reduced if trading programs for 

point and non-point sources are properly developed and implemented concurrently with 

regulations.

Protecting and Enhancing Social Welfare in the Basin 

 Implementing the management options needed to reduce nutrients will clearly 

affect the social welfare of many who live in the basin. On the positive side, N and P 

reductions will improve environmental quality within the basin and, as the Integrated 

Assessment documented, these co-benefits can be highly valuable. Second, if the costs of 

implementing these management options are borne largely by residents in the region, 

then preserving/enhancing social welfare will require implementing policies that target 

the most cost-effective sources and locations for nutrient reductions. 

 Subsidies, not regulation, have been the government’s primary tool for managing 

agricultural production and income support in the U.S., as well as conservation in 

agriculture. Hence re-structuring subsidies and conservation programs represents an 

important tool for reducing nutrient runoff from agricultural production. The Integrated 

Assessment recognized numerous agricultural management practices that improve water 

quality but did not discuss the efficiency of the tools for their implementation. A large 

body of economics literature exists regarding the relative merits and cost-effectiveness of 

taxes, regulations, voluntary approaches, permit trading, subsidies, and other instruments 

that could apply to reducing nutrient losses. This research indicates that if significant 

behavioral changes are to be realized, incentives are needed across a wide range of 

sectors. Such incentives can be positive (e.g., subsidies) or negative (e.g., taxes or 

direction regulation with enforcement actions), but they must be strong enough to change 

behavior. A thorough and quantitative comparison of all possible incentives for all 

sectors was beyond the SAB Panel’s scope; however, research indicates that the 

following approaches are cost-effective. 

 First, the establishment (and continuation where appropriate) of targeting and 

competitive bidding mechanisms results in lands enrolled in conservation programs (e.g., 

the Conservation Reserve Program, the Environmental Quality Incentives Program, and 

the Conservation Security Program) that achieve maximum environmental benefits. 

Moreover, conservation compliance requirements extended to nutrient management, if 

adequately monitored and enforced, could be cost-effective. Targeting conservation 

practices to the locations within a watershed where they produce the most N and P 
9

reductions (and co-benefits) and targeting entire watersheds that have relatively high N 

and/or high P contributions are both cost-effective targeting approaches. 

 Second, economic incentives are needed for the full range of conservation 

options. Incentives for development of technologies to convert cellulosic perennials to 

biofuels would be needed to greatly reduce N and P losses from agricultural systems. Restructuring eligibility requirements for existing subsidies to reward conservation in all its 

forms (in-field nutrient management, cover crops, conservation buffers, wetlands, 

alternative drainage, manure management) could help mitigate the unintended 

consequences of agricultural production. 

Conclusion 

 In sum, environmental decisions and improvements require a balance between 

research, monitoring and action. In the Gulf of Mexico, the action component lags 

behind the growing body of science. Moreover, certain aspects of current agricultural 

and energy policies conflict with measures needed for hypoxia reduction. Although 

uncertainty remains, there is an abundance of information on how to reduce hypoxia in 

the Gulf of Mexico and to improve water quality in the MARB, much of it highlighted in 

the Integrated Assessment. To utilize that information, it may be necessary to confront 

the conflicts between certain aspects of current agricultural and energy policies on the 

one hand and the goals of hypoxia reduction and improving water quality on the other. 

This dilemma is particularly relevant with respect to those policies that create economic 

incentives. The SAB Panel's recommendation to address the structure of economic 

incentives stems from sound science. 

 Basing management decisions on sound science means taking action at several 

different scales, addressing conflicts between policies, and acting in the face of 

uncertainties. Lessons learned from current actions can inform and improve future 

decisions. While actions must come first, they must also be coupled with monitoring and 

modeling of management activities within a conceptual framework to improve 

understanding of the system. Done well, this process of adaptive management means 

that, over time, society will benefit from cost-effective environmental decisions that 

reduce hypoxia in the Gulf and improve water quality in the MARB. 
10

1. Introduction 

1.1. Hypoxia and the Northern Gulf of Mexico – A Brief Overview 

 Nutrient over enrichment from anthropogenic sources is a major stressor of 

aquatic, estuarine, and marine ecosystems. Nutrients enter ecosystems through off-target 

migration of fertilizer from agricultural fields, golf courses, and lawns; disposal of animal 

manure; atmospheric deposition of nitrogen; erosion of soil containing nutrients; sewage 

treatment plant discharges; and other industrial discharges. Excessive nutrients promote 

nuisance blooms (excessive growth) of opportunistic bacteria, cyanobacteria, and algae. 

When the available nutrients in the water column have been sequestered in plant biomass, 

the nuisance blooms die, decompose, and deplete dissolved oxygen in the water column 

and at the sediment water interface. This oxygen depletion, known as hypoxia, occurs 

when normal dissolved oxygen concentrations in shallow coastal and estuarine systems 

decrease below the level required to support many estuarine and marine organisms (≤ 2 

mg/L). 

 Hypoxia can occur naturally in deep basins, fjords, and oxygen minimal coastal 

zones associated with upwelling. However, nutrient induced hypoxia in shallow coastal 

and estuarine systems is increasing worldwide. A large hypoxic area, averaging about 

16,500 km2

 (10, 250 mi2) and ranging from 8,500 to 22,000 km2 (3,100 to 7,700 mi2) 

forms annually between May and September in the northern Gulf of Mexico. Shown in 

Figure 1, the northern Gulf hypoxic zone is the largest in the United States and the 

second largest worldwide. Hypoxic conditions result from complex interactions between 

climate, weather, basin morphology, circulation patterns, water retention times, 

freshwater inflows, stratification, mixing, and nutrient loadings. Nutrient fluxes from the 

Mississippi-Atchafalaya River basin (MARB), coupled with temperature and density 

induced stratification have been implicated as the primary cause of hypoxia in the 

northern Gulf of Mexico (NGOM) (CENR, 2000). 

Figure 1: Map of the frequency of hypoxia in the northern Gulf of Mexico, 1985-2005. Taken from N.N. 

Rabalais, LUMCON, 2006. 
11

 The MARB is one of the largest river systems in the world (Figure 2), draining 

approximately 40% of the contiguous United States, and is the largest contributor of 

freshwater and nutrients to the NGOM. About two thirds of the total Mississippi River 

flow enters the northern Gulf via the Mississippi River delta. The remaining third is 

diverted to the Atchafalaya River and eventually enters the northern Gulf about 200 km 

west of the main Mississippi River delta. Prevailing east-to-west currents in the Gulf 

move much of the freshwater, suspended sediments, and dissolved and particulate 

nutrients onto the Louisiana-Texas continental shelf. 

Figure 2: Map showing the extent of the Mississippi-Atchafalaya River basin. 

 Land-use activities in the MARB influence water quality in the entire watershed 

as well as in the NGOM. Low oxygen events on the Louisiana-Texas continental shelf 

have been reconstructed over the past 180 years using the relative abundance of lowoxygen-tolerant benthic foraminifera in sediment cores (Osterman et al., 2005). These 

data show that the prevalence of low oxygen events has increased over the past 50 years. 

Several hypoxic events from 1870 and 1910 (prior to widespread fertilizer use) were 

attributed to natural variation in river flow that enhanced freshwater and nutrient 

transport. The increased prevalence over the past several decades is clearly related to 

increased nutrient loads. However, there is substantial variation in year-to-year inputs of 

both freshwater and nutrients from the MARB. Since these are correlated, it is not 

possible to tease apart the relative importance of increased eutrophication versus 

increased stratification in any given year over the recent past. Clearly, land-use practices 

in the MARB affect watershed dynamics and water quality within the Basin as well as the 
12

northern Gulf. Land-use practices in the Basin are also influenced by various, and 

conflicting, national environmental, conservation and agricultural policies. 

1.2. Science and Management Goals for Reducing Hypoxia 

 In 1997, the U. S. Environmental Protection Agency (EPA) established the 

Mississippi River/Gulf of Mexico Watershed Nutrient Task Force (or Task Force). The 

Task Force brought together federal agencies, states and tribes to consider options for 

reducing, mitigating, and controlling hypoxia in the NGOM. The Task Force requested 

that the White House National Science and Technology Council (NSTC) conduct a 

scientific assessment of the causes and consequences of Gulf hypoxia. The NSTC 

Committee on Environment and Natural Resources (CENR) formed a federal intraagency Hypoxia Working Group to plan and conduct the assessment. The need for the 

assessment was given additional impetus by passage of the Harmful Algal Bloom and 

Hypoxia Research and Control Act of 1998. The Act specifically called for an integrated 

scientific assessment of causes and consequences of hypoxia in the Gulf of Mexico and a 

plan of action to reduce, mitigate, and control hypoxia. 

The scientific assessment was led by the National Oceanic and Atmospheric 

Administration (NOAA) with oversight among several federal agencies. As a first step, 

six reports (available at http.//www.nos.noaa.gov/products/pub_hypox.html) covering 

key topics were developed. These include characterization of hypoxia (Rabalais et al., 

1999a); ecological and economic consequences of hypoxia (Diaz and Solow, 1999); flux 

and sources of nutrients in the Mississippi-Atchafalaya River basin (Goolsby et al., 

1999); effects of reducing nutrient loads to surface waters within the Mississippi River 

basin and Gulf of Mexico (Brezonik et al., 1999); reducing nutrient fluxes, especially 

nitrate-nitrogen, to surface water, ground water, and the Gulf of Mexico (Mitsch et al., 

1999); and evaluation of the economic costs and benefits of the methods for reducing 

nutrient fluxes to the Gulf of Mexico (Doering et al., 1999). 

 The six NOAA reports provided the scientific foundation for the Integrated 

Assessment of Hypoxia in the Northern Gulf of Mexico (CENR, 2000) (or Integrated 

Assessment, available at http://oceanservice.noaa.gov/products/pubs_hypox.html). The 

Integrated Assessment concluded that hypoxia in the northern Gulf was caused by excess 

nitrogen from the MARB, in combination with stratification of Gulf waters. Informed by 

the Integrated Assessment, in 2001 the Task Force completed its Action Plan for 

Reducing, Mitigating and Controlling Hypoxia in the Northern Gulf of Mexico 

(MR/GMWNTF, 2001) (or Action Plan, available at 

http://www.epa.gov/msbasin/taskforce/actionplan.htm). The Action Plan described three 

primary hypoxia management goals. 

1. Coastal Goal: By the year 2015, subject to the availability of additional 

resources, reduce the five-year running average of the areal extent of the 

Gulf of Mexico hypoxic zone to less than 5,000 km2

 (1,930 mi2) through 

implementation of specific, practical, and cost-effective voluntary actions 

by all states, tribes, and all categories of sources and removals within the 
13

Mississippi-Atchafalaya River basin to reduce the annual discharge of 

nitrogen into the Gulf. 

2. Within Basin Goal: To restore and protect the waters of the 31 states and 

tribes within the MARB through implementation of nutrient- and 

sediment- reduction actions to protect public health and aquatic life as 

well as reduce negative impacts of water pollution on the Gulf of Mexico. 

3. Quality of Life Goal: To improve the communities and economic 

conditions across the Mississippi-Atchafalaya River basin, in particular 

the agriculture, fisheries, and recreation sectors, through improved public 

and private land management and a cooperative incentive based approach. 

In 2005, the Task Force recognized a need to update the Integrated Assessment 

and Action Plan with more recent science. Accordingly, the Task Force sponsored four 

symposia on the upper Mississippi River basin; Gulf Hypoxia; the lower Mississippi 

River basin, and Nutrient Sources, Fate and Transport. Each of the symposia focused on 

scientific developments since 1999. In conjunction with the symposia, the Task Force 

also developed a bibliography of recent literature on hypoxia causes, effects, and control 

options since the year 2000 (available at 

http://www.epa.gov/msbasin/taskforce/reassess2005.htm). In addition to science 

activities, the Task Force also compiled information necessary for nutrient management 

and control in the MARB in two reports. The Management Action Review Team Report 

(MART, 2006a) summarized federal programs that encouraged watershed planning and 

land-use practices to reduce nutrient loadings. The Reassessment of Point Source 

Nutrient Mass Loadings to the Mississippi River Basin report (MART, 2006b) updated 

annual mass loading estimates for total nitrogen (TN), total phosphorus (TP), and 

biochemical oxygen demand (BOD) (Task Force documents are available at 

http://www.epa.gov/msbasin/taskforce/reassess2005.htm.) The Task Force is also 

working with the U. S. Department of Agriculture’s (USDA) Conservation Effects 

Assessment Program (CEAP) to encourage the quantification and documentation of 

environmental effects and benefits of conservation practices on agricultural lands to 

control nutrients in the MARB. CEAP documents are available at 

http://www.nrcs.usda.gov/Technical/nri/ceap/. 
14

1.3. EPA Science Advisory Board (SAB) Hypoxia Advisory Panel 

On behalf of the Task Force, EPA’s Office of Water requested that the Science 

Advisory Board (SAB) evaluate the state-of-the-science regarding hypoxia in the Gulf of 

Mexico and potential nutrient mitigation and control options in the MississippiAtchafalaya River basin. In response to this request, the SAB established the SAB 

Hypoxia Advisory Panel (SAB Panel). The Office of Water asked the SAB Panel to 

focus its evaluation on the following issues and questions. 

1. Characterization of Hypoxia – The development, persistence and areal extent 

of hypoxia is thought to result from interactions in physical, chemical and 

biological oceanographic processes along the northern Gulf continental shelf; 

and changes in the Mississippi River basin that affect nutrient loads and fresh 

water flow. 

A. Address the state-of-the-science and the importance of various processes 

in the formation of hypoxia in the Gulf of Mexico. These issues include: 

i. increased volume or funneling of fresh water discharges from the 

Mississippi River; 

ii. changes in hydrologic or geomorphic processes in the Gulf of Mexico 

and the Mississippi River basin; 

iii. increased nutrient loads due to coastal wetlands losses, upwelling or 

increased loadings from the Mississippi River basin; 

iv. increased stratification, and seasonal changes in magnitude and 

spatial distribution of stratification and nutrient concentrations in the 

Gulf; 

v. temporal and spatial changes in nutrient limitation or co-limitation, for 

nitrogen or phosphorus, as significant factors in the development of the 

hypoxic zone;and 

vi. the implications of reduction of phosphorus or nitrogen without 

concomitant reduction of the other. 

B. Comment on the state of the science for characterizing the onset, volume, 

extent and duration of the hypoxic zone. 

2. Characterization of Nutrient Fate, Transport and Sources -- Nutrient loads, 

concentrations, speciation, seasonality and biogeochemical recycling processes 

have been suggested as important causal factors in the development and 

persistence of hypoxia in the Gulf. The Integrated Assessment (CENR 2000) 

presented information on the geographic locations of nutrient loads to the Gulf 

and the human and natural activities that contribute nutrient loadings. 

A. Given the available literature and information (especially since 2000), 

data and models on the loads, fate and transport and effects of nutrients, 

evaluate the importance of various processes in nutrient delivery and effects. 

These may include: 
15

i. the pertinent temporal (annual and seasonal) characteristics of nutrient 

loads/fluxes throughout the Mississippi River basin and, ultimately, to the 

Gulf of Mexico; 

ii. the ability to determine an accurate mass balance of the nutrient loads 

throughout the basin;and 

iii. nutrient transport processes (fate/transport, sources/sinks, 

transformations, etc.) through the basin, the deltaic zone, and into the 

Gulf. 

B. Given the available literature and information (especially since 2000) on 

nutrient sources and delivery within and from the basin, evaluate capabilities 

to: 

i. predict nutrient delivery to the Gulf, using currently available scientific 

tools and models; and 

ii. route nutrients from their various sources and account for the transport 

processes throughout the basin and deltaic zone, using currently available 

scientific tools and models. 

3. Scientific Basis for Goals and Management Options -- The Task Force has 

stated goals of reducing the 5-year running average areal extent of the Gulf of 

Mexico hypoxic zone to less than 5,000 square kilometers by the year 2015, 

improving water quality within the basin and protecting the communities and 

economic conditions within the basin. Additionally, nutrient loads from various 

sources in the Mississippi River basin have been suggested as the major driver for 

the formation, extent and duration of the Gulf hypoxic zone. 

A. Are these goals supported by present scientific knowledge and 

understanding of the hypoxic zone, nutrient loads, fate and transport, sources 

and control options? 

i. Based on the current state-of- the-science, should the reduction goal for 

the size of the hypoxia zone be revised? 

ii. Based on the current state-of-the-science, can the areal extent of Gulf 

hypoxia be reduced while also protecting water quality and social welfare 

in the basin? 

B. Based on the current state-of- the-science, what level of reduction in 

causal agents (nutrients/discharge) will be needed to achieve the current 

reduction goal for the size of the hypoxic zone? 

C. Given the available literature and information (especially since 2000) on 

technologies and practices to reduce nutrient loss from agriculture, runoff 

from other non-point sources and point source discharges, discuss options 

(and combinations of options) for reducing nutrient flux in terms of cost, 

feasibility and any other social welfare considerations. These options may 

include: 
16

i. the most effective agricultural practices, considering maintenance of 

soil sustainability and avoiding unintended negative environmental 

consequences; 

ii. the most effective actions for other non-point sources; and 

iii. the most effective technologies for industrial and municipal point 

sources. 

In all three areas, please address research and information gaps (expanded 

monitoring, documentation of sources and management practices, effects of 

practices, further model development and validation, etc.) that should be 

addressed prior to the next 5-year review.

1.4. The SAB Panel’s Approach 

The NOAA, CENR, and Task Force documents (see Section 1.2 above) provide a 

comprehensive scientific review of hypoxia causes, and potential mitigation and control 

actions through about 1999 to 2000. Further, more recent science and management 

information on the Gulf and MARB has been captured in the Task Force sponsored 

symposia, literature search, MART reports, and CEAP activities. Accordingly, the SAB 

Panel initiated its deliberations by reviewing these documents. The SAB Panel invited 

the chairs of the four symposia to present summaries of key findings, and also invited 

selected researchers (see acknowledgements) currently working on hypoxia issues to 

present their recent work. The SAB Panel also relied on the individual and collective 

experience and expertise of its members to provide additional relevant publications and 

information to assist its deliberations. The SAB Panel convened four public face-to-face 

meetings and 15 public teleconferences to deliberate and develop this state-of-the-science 

report (background and other materials for the meetings may be found at: 

http://www.epa.gov/sab/panels/hypoxia_adv_panel.htm). 

The SAB Panel recognized the inherent complexity and connectivity between the 

Mississippi –Atchafalaya River basin and Gulf of Mexico and agreed that a systems 

perspective within an adaptive management framework was needed. The systems 

approach allowed understanding of feedback loops so that perturbations in one part of a 

system affect the interrelationships and stability of the system as a whole. Adaptive 

management seeks to maximize flexibility in management so that learning and 

adjustments can occur. Adaptive management employs six basic operating principles: 1) 

resources of concern are clearly defined; 2) conceptual models are developed during 

planning and assessment; 3) management questions are formulated as testable hypotheses 

to guide inquiry; 4) management actions are treated like experiments that test hypotheses 

to answer questions and provide future management guidance; 5) ongoing monitoring 

and evaluation is necessary to improve accuracy and completeness of knowledge; and 6) 

management actions are revised with new cycles of learning. 

This report considers models as essential for understanding the inherent 

complexities of the MARB and the NGOM. Additionally, the collection of critical data 

at appropriate spatial and temporal scales is absolutely necessary to optimize future 
17

research and management actions. Data collection should be based on a well-defined 

conceptual model of the overall system. Monitoring programs will often provide data for 

existing models and assist with broader interpretations of data and information. In 

summary, a systems perspective combined with an adaptive management approach will 

greatly enhance scientific understanding and management of hypoxia in the MARB and 

the NGOM. 

This report deals largely with the review of research and findings since the 

Integrated Assessment. Background material and findings prior to 2000 are used when 

appropriate or when instrumental to understanding the relative importance of more recent 

work. However, those interested in the details of the Integrated Assessment and the six 

topical reports that provided the scientific basis for the assessment are referred directly to 

those documents. 
18

2. Characterization of Hypoxia 

2.1. Processes in the Formation of Hypoxia in the Gulf of Mexico. 

The hypoxic region along the northern Gulf of Mexico (NGOM) extends up to 

125 km offshore and to 60 m water depth, has substantial variability with an average midsummer areal extent of 16,500 km2

 (2001-2007), and extends in some years from the 

Mississippi River mouth westward to Texas coastal waters (Rabalais et al., 2007). This 

hypoxic region (Figure 1) occurs along a relatively shallow, open coastline with complex 

circulation and water column structure typical of many coastal regions and includes 

massive inputs of freshwater, weak tidal energies, seasonally varying stratification 

strength, generally high water temperature, wind effects from both frontal weather 

systems and hurricanes, and mixing of river plumes from the Atchafalaya and Mississippi 

Rivers and other smaller sources (DiMarco et al., 2006; Hetland and DiMarco, 2007). 

The plumes of the Mississippi and Atchafalya Rivers can be observed as areas of highly 

turbid low salinity surface water. The limits of these plumes have been defined in 

different ways, but in satellite imagery their boundaries can be clearly observed as sharp 

color discontinuities. Since the release of the Integrated Assessment and the Action Plan

in 2001, the measured areal extent of the hypoxic region has averaged 16,500 km2

, with a 

range of 8,500 to 22,000 km2

. Many reports from both the Integrated Assessment and 

post-Integrated Assessment periods concluded that physical and morphological 

characteristics such as these make the NGOM prone to hypoxic conditions. 

2.1.1. Historical Patterns and Evidence for Hypoxia on the Shelf 

An important question regarding hypoxia on the Mississippi River shelf is how far 

back in time has hypoxia been observed? Is it a recent phenomenon or has hypoxia been 

a regular natural feature of a productive shelf region? Unfortunately the monitoring data 

are not entirely sufficient to address this question, for only a limited number of 

measurements are available prior to the time when wide-spread hypoxia was first 

observed on the Louisiana shelf in the mid-1980s (Rabalais et al., 1999a). However, a 

limited number of additional paleoecological studies have been carried out on the 

Mississippi River shelf since the Integrated Assessment. All studies from dated sediment 

cores show recent increases in low oxygen concentrations with time, although the precise 

timing and response varies depending upon the proxy studied and the dating of cores. 

The accumulated body of evidence shows that the pattern of change is concomitant with 

recent (since the 1960s) increases in nutrient loading from the Mississippi River causing 

increasingly severe hypoxia on the shelf. The spatial distribution of reliably dated 

sediment cores, with most cores taken on the southeastern Louisiana shelf just west of the 

Mississippi River delta, is not sufficient to determine the increases in the spatial extent of 

hypoxia with time. 

 A limiting factor in all paleoecological studies is the availability of undisturbed 

sediment cores to provide an accurate picture of changes through time. This is a 

particular challenge in a hydrologically dynamic, relatively shallow environment as 

found on the Mississippi River shelf with resuspension processes, movement of fluid 
19

muds, mixing by benthic organisms, and more recently sediment disturbance of upper 

sediment layers through bottom trawling. Despite these challenges, a number of 

reasonably dated sediment cores, primarily within the Louisiana bight, have provided a 

coherent picture of changes in hypoxia with time. 

 Bacterial pigments measured in sediments at one location on the Louisiana shelf 

were characteristic of anoxygenic phototrophic sulfur bacteria and have their highest 

concentrations between 1960 and the present (Chen et al., 2001). These bacteriopigments 

were not present prior to 1900. Further evidence of increased hypoxia is provided by 

Chen et al. (2001) using algal pigments, which show increases in the 1960s. The increase 

in these pigments reflects enhanced preservation with hypoxia as well as nutrient-driven 

increases in production. Rabalais et al. (2004, 2007) also report increases in algal 

pigment concentrations over time from a number of sediment cores, with gradual changes 

from 1955 to 1970, followed by a steady increase to the late 1990s. However, the 

patterns observed by Rabalais et al. (2004, 2007) are confounded by the rapid 

degradation of carbon and algal pigments in upper surface sediments with most studies of 

sediment pigments correcting for diagenesis by normalizing pigments with organic 

carbon (Leavitt and Hodson, 2001). In addition, there is some evidence for spatial 

increases in hypoxic extent through time: increases in pigment concentrations from one 

sediment core from west of the Atchafalaya River outflow suggests that nutrient-driven 

increases in production occurred later at this location than in the Mississippi River Bight 

(Rabalais et al., 2004). There has been an increased accumulation of total organic carbon 

and biogenic silica in recent sediments near the mouth of the Mississippi River (Turner 

and Rabalais, 1994; Turner et al., 2004), although the spatial and temporal variations 

observed between dated sediment cores are large. 

 Several studies have examined changes in the benthic foraminiferal community in 

dated sediment cores (Platon and Sen Gupta, 2001; Osterman et al., 2005; Platon et al., 

2005). Different species of bottom living benthic foraminifera are particularly sensitive 

to changes in bottom water oxygen concentrations, and the abundance of these species is 

a widely used indicator of hypoxia. Significant changes in the composition of the benthic 

foraminiferal community have occurred in the past century. Several indicators, e.g., the 

PEB index (the relative abundance of three low-oxygen tolerant species of benthic 

foraminifers; Pseudononin altlanticum, Epistominella vitrea, and Buliminella morgani) 

(Osterman et al., 2005) and the A/P ratio (agglutinated to porcelaneous orders) (Platon et 

al., 2005) indicate that increases in the occurrence of low oxygen events have occurred 

over the past 50 years (Figure 3). In addition, the porcelaneous genus Quinqueloculina, 

an organism that occurs where dissolved oxygen concentrations are higher than 2 mg/l, 

was present but has disappeared from the foraminiferal community since 1900, indicating 

that prior to this time there was sufficient oxygen at the sediment-water interface to 

enable survival of such species (Rabalais et al., 2007). Osterman et al. (2005) have 

shown that several probable low oxygen events that occurred in the past 180 years are 

associated with high Mississippi River discharge rates, although the recent changes in 

foraminiferal communities are more extreme than any that occurred in the past. The data 

support the interpretation that hypoxia is a recent phenomenon and has been amplified 

from an otherwise naturally occurring process. 
20

Figure 3: Plots of the PEB index (%PEB) in sediment cores from the Louisiana shelf. Higher values of the 

PEB index indicate lower dissolved oxygen contents in bottom waters. Taken from Osterman et al. (2005). 

Key Findings and Recommendations 

 The SAB Panel finds that the paleoecological data are consistent with increased 

prevalence of hypoxic conditions in recent decades. However, the spatial distribution of 

sediment cores is not sufficient to determine the increases in the spatial extent of hypoxia 

with time. Although given the complex nature of disturbance, there may be limited 

opportunities to determine temporal changes in the extent of hypoxia. To advance the 

understanding of spatial and temporal trends in hypoxia in the NGOM, the SAB Panel 

offers the following recommendations. 

• In future research on the Mississippi River shelf, more attention should be 

focused on establishing reliable chronologies in additional sediment cores. 

• In order to establish spatial changes in hypoxia over time, where possible 

additional sediment cores should be collected over a broader area of the 

Mississippi River shelf. 
21

2.1.2. The Physical Context 

Oxygen budget: general considerations

 The oxygen budget on the NGOM shelf is influenced by several sink and source 

terms. Oxygen (O2) concentration in the bottom layer will decrease and possibly become 

hypoxic or even anoxic when the export and consumption of oxygen by respiration 

exceed the import or production of “new” oxygenated water by photosynthesis. 

Mathematically, this relationship can be expressed in its simplest form by the following 

oxygen balance equation: 

 (1) 

in which the left-hand term represents the change of oxygen concentration with time; 

term (1) on the right represents the horizontal advection by across-shelf currents, u; term 

(2) represents the horizontal advection by along-shelf currents, v; term (3) represents 

vertical transport by upwelling or downwelling; term (4) represents vertical mixing and 

Kz (x,y,z) is the vertical eddy diffusivity; term (5) represents horizontal diffusion and 

KH (x,y,z) is the horizontal eddy diffusivity; term (6) is oxygen flux across the air-sea 

interface; term (7) is the non-conservative sink (i.e., oxygen consumption); and term (8) 

refers to in situ production of oxygen by photosynthesis. The horizontal advection terms 

may reflect contributions from tides, wind stress, buoyancy, and momentum input from 

rivers, large-scale and mesoscale eddies, or topographically trapped shelf waves. Threedimensional hydrodynamic models are required to adequately account for these 

contributions (Morey et al., 2003a, 2003b; Hetland and DiMarco, 2007). The respiration 

term (7) relates directly to organic matter mineralization and must be understood in the 

context of water column and sediment biogeochemical processes described in later 

sections. As depicted in equation 1, the change in oxygen concentration with time at any 

point in the water column is affected by sources and sinks of oxygen at and below the 

surface. Term 6 (oxygen flux across the air-sea interface) represents a surface source and 

sink, while term 8 (photosynthesis) is a source of oxygen in waters beneath the air-sea 

interface. Although equation 1 above suggests that alongshore and cross-shore dispersion 

coefficients are of equal magnitude, the Panel notes that this has not been demonstrated. 

The effects of cross-shore dispersion processes must be parameterized and additional 

research on lateral mixing processes must be completed before such parameterization can 

be performed with confidence. 

Vertical mixing as a function of stratification and vertical shear 

 Over the Louisiana-Texas shelf, the vertical mixing term (4) plays a key role in 

the local oxygen balance. Its magnitude depends on the value of vertical eddy diffusivity 

Kz, which is highly variable in both space and time and depends on the gradient 

Richardson number Ri (MacKinnon and Gregg, 2005), defined by 

2 22

2 2 2 2 2 22

z H as 2 22

O O O O O OO u v w K K F Resp. photosynthesis t x y z z xy

Change 1) (2) (3) (4) (5) (6) (7)

∂ ∂ ∂ ∂ ∂ ∂∂ ⎛ ⎞

= − − − + + + + − + ⎜ ⎟ ∂ ∂ ∂ ∂ ∂ ∂∂ ⎝ ⎠

 ( 

r

 (8)
22

2 2 2

2

⎟

⎠

⎞ ⎜

⎝

⎛

∂

∂

⎟

⎠

⎞ ⎜

⎝

⎛

∂

− ∂

=

⎟

⎠

⎞ ⎜

⎝

⎛

∂

∂ ⎟ + ⎠

⎞ ⎜

⎝

⎛

∂

∂ =

z

V

z

g

z

v

z

u

N Ri

ρ

ρ (2) 

where N is an index of stratification strength known as the buoyancy frequency, ρ is the 

water density, g is the gravitational acceleration (9.8 m/s2

), and ∂V/∂z is the vertical shear 

of horizontal current. The gradient Richardson number, Ri, expresses the ratio of 

turbulence suppression by stratification (numerator) relative to vertical shear production 

of turbulence (denominator). When Ri > 1⁄4, turbulence is suppressed, and vertical 

transport of oxygen from surface to bottom layers by turbulent mixing is unlikely to 

occur. Thus, strong vertical density gradients (for example, when freshwater sits on top 

of salty water) and/or weak current shears can suppress vertical mixing and be favorable 

to hypoxia. Key physical factors that produce stronger vertical density gradients (∂ρ/∂z) 

and thus reduce vertical mixing include freshwater inputs from rivers or precipitation, 

warmer surface temperatures from absorption of solar radiation or sensible heat input, 

and near-bed suspended sediment (which causes benthic stratification). Factors 

responsible for producing enhanced vertical shear (∂V/∂z) and enhanced vertical mixing 

include tidal and wind-driven currents, inertial waves, internal tides, surface waves and 

Langmuir cells (Kantha and Clayson, 2000). Although no field studies of vertical mixing 

by microstructure measurements of the turbulent dissipation rates of velocity, salinity and 

temperature fluctuations have been reported for the NGOM, many of the physical 

mechanisms described on the New England shelf (MacKinnon and Gregg, 2005) and in 

Monterrey Bay (Carter et al., 2005) are at play on the NGOM as well. 

 While the tributaries within the Mississippi River basin are the sources of nutrient 

loading to the river trunk, the distributaries within the Mississippi Delta are critical to the 

final dispersal of nutrients, buoyancy and sediment into the Gulf of Mexico. The 

multiple distributary mouths of the Mississippi and Atchafalaya Rivers are, for the most 

part, highly stratified “salt wedge” estuaries, and their combined effluent debouches onto 

the shelf as a discrete layer of fresh water that is spread into the surface layer. Exceptions 

occur where smaller distributaries enter shallow bays where salinity is nearly uniform 

from top to bottom. Total buoyancy fluxes are, of course, proportional to river discharge 

and cause the turbulence suppressing stratification of the upper water column that is 

strongly implicated in hypoxia. In most inner shelf environments, tidal currents are the 

major source of mixing, and the position of temperature fronts (sharp horizontal 

temperature gradients) can often be accurately predicted from the h/Ut

3

 criterion of 

Simpson and Hunter (1974), where h is the local depth and Ut represents the depthaveraged tidal velocity. Unfortunately, the Simpson-Hunter criterion of tidal mixing has 

not yet been mapped for the northern Gulf of Mexico. Nevertheless, it is generally 

agreed that tidal mixing over the Louisiana-Texas shelf is very weak because the tidal 

range is only about 40 cm and tidal currents typically do not exceed 10 cm/s (Kantha, 

2005). So the contribution of tidal mixing to the vertical exchange of oxygen is minimal 

over the shelf, particularly off the mouths of the larger distributaries, such as Southwest 
23

and South Passes, which debouch into deep water. Wind-driven currents are stronger 

than tidal currents but occur episodically (Ohlmann and Niiler, 2005). Winds also cause 

breaking and white capping waves as well as vertical circulation (Langmuir) cells 

(Thorpe, 2004) that contribute to mixing in the upper water column. 

 The hydrologic regime of the Mississippi River and the spatial distribution and 

timing of freshwater inputs to the shelf relative to the occurrence of energetic currents 

and waves are critical to vertical mixing intensity, stratification, and hypoxia. These 

influences were recognized in the CENR report (Rabalais et al., 1999). Using oxygen 

measurements within 2 m of the bottom and vertical profiles of temperature and salinity 

collected during the 1992-1994 LaTex experiment on the Louisiana-Texas shelf and 

during the 1996-1998 NECOP (Northe36astern Gulf of Mexico Chemical Oceanography 

Program) in the region east of the Mississippi delta and north of Tampa Bay, Belabassi 

(2006) performed an evaluation of the empirical relationships between the maximum 

value of the buoyancy frequency Nmax in the water column, bottom silicate concentration 

as a proxy of phytoplankton remineralization, and the occurrence of hypoxic waters (< 2 

mg/L) or low-oxygen waters (< 3.4 mg/L). She found that low-oxygen and hypoxic 

bottom waters only occurred when Nmax, evaluated at a vertical resolution of 0.5 m was 

greater than 40 cycles per hour (cph), which corresponds to a buoyancy period shorter 

than 1.5 minutes. This result confirms that strong density stratification is a prerequisite 

for hypoxia occurrence on the northern Gulf of Mexico shelf. She also found that lowsalinity water from the Mississippi and Atchalafaya Rivers was generally the main 

contributor to stratification in spring and summer, although temperature was more 

important than salinity in determining stratification during summer at all depths west of 

Galveston Bay and at depths greater than 20 m between Galveston Bay and Terrebonne 

Bay. Interestingly, stations with strong stratification (Nmax greater than 40 cph) but low 

bottom silicate concentrations (less than 18 mmol m-3) did not have low-oxygen or 

hypoxic bottom waters. The analyses of Belabassi (2006) thus indicate that strong 

stratification (Nmax greater than 40 cph) is a necessary but not sufficient condition for 

bottom layer hypoxia; a second necessary condition for hypoxia occurrence is high 

bottom water remineralisation as indicated by the proxy of high concentrations of bottom 

water silicates (greater than 18 mmol m-3). Simply put, there cannot be hypoxia without 

both density stratification and degradation of labile organic matter. 

 Stow et al. (2005) attempted to disentangle the relative contributions of 

eutrophication and stratification as drivers of hypoxia in the NGOM. Their analysis 

indicates that the probability of observing bottom hypoxia increases rapidly when the top 

to bottom salinity difference reaches a threshold of 4.1. Stow et al. (2005) also showed 

that this salinity threshold decreased from 1982 to 2002. Concurrently, they highlighted 

that surface temperature had increased, while surface dissolved oxygen decreased, 

suggesting that changes in surface mixed layer properties may be partly responsible for 

oxygen decrease in the bottom layer. 
24

Changes in Mississippi River hydrology and their effects on vertical mixing 

 By far the most important change in local hydrology has been the increased flow 

of the Atchafalaya River during the 20th century. Available data show that in the early 

1900’s the discharge from the Atchafalaya River accounted for less than 15% of the 

combined Atchafalaya-Mississippi River discharge (Figure 4). This proportion 

progressively increased to reach about 30% in 1960, peaked at 35% in 1975 and since 

then was reduced to 30% by means of regulatory measures (Bratkovich et al., 1994). To 

understand the significance of this change on circulation patterns and on the strength of 

stratification on the Louisiana-Texas shelf, it must be kept in mind that the Mississippi 

River plume enters the shelf near the shelf edge and typically does not extend to the 

bottom, even near the river mouth. On the other hand, the Atchafalaya River plume 

enters a broader shelf, is more diffuse, and extends to the bottom over a larger distance 

from the river mouth. 

Figure 4: Change in the relative importance of the Atchafalaya flow to the combined flows from the 

Mississippi and Atchafalaya Rivers over the 20th Century. Reprinted from Bratkovich et al. (1994). 

 The short distances (10 to 30 km) separating Mississippi River delta passes from 

the shelf break facilitate the export of plume waters offshore and to the east by sporadic 

wind events or by eddies present on the upper continental slope, some of which may have 

been spun off by the Loop Current (Ohlmann and Niiler, 2005; Oey et al., 2005a, 2005b). 

The modeling study of Morey et al. (2003a) shows that a prime export pathway for river 

freshwater during the summer months is to the east, and offshore of the Mississippi River 

delta. During non-summer months, the main freshwater export pathway consists of a 

coastal jet flowing westward to Texas and then southward. Etter et al. (2004) estimate 

that 43% ± 10% of the Mississippi River discharge is carried westward to the LouisianaTexas continental shelf, the remainder being carried offshore and/or eastward. While this 

proportion is slightly lower than the earlier estimate of 53% ± 10% from Dinnel and 

Wiseman (1986), both studies indicate that roughly half of the freshwater from the 

Mississippi River goes westward, toward the Louisiana-Texas continental shelf. 

 In contrast, 100% of the Atchafalaya River discharge of freshwater, nutrients and 

sediments is delivered to the Louisiana-Texas continental shelf. Moreover, the very 

broad shelf near Atchafalaya Bay implies longer residence times of this freshwater source 

on the shelf compared with freshwater from the Mississippi River delta. A “back-of-theenvelope” calculation helps capture the full significance of the increased Atchafalaya 
25

River flow. In the early 1900’s, for every 100 m3

 of water discharged, 85 m3 took the 

Mississippi River delta route. Of these, roughly 42.5 m3

 went westward and 42.5 m3

went offshore or eastward. The 42.5 m3

 that went westward were added to the 15 m3 that 

took the Atchafalaya River route to give a grand total of 57.5 m3 of freshwater on the 

Louisiana-Texas continental shelf. By contrast, in the post-1970’s, for every 100 m3

 of 

combined Atchafalaya and Mississippi River outflows, 70 m3

 took the Mississippi River 

route. Of these, roughly 35 m3

 went westward, and 35 m3 went offshore or eastward. 

The 35 m3

 that went westward were added to the 30 m3 that took the Atchafalaya River 

route to give a grand total of 65 m3 of freshwater on the Louisiana-Texas continental 

shelf. This simple calculation reveals two things. First, it suggests that even in the 

absence of a temporal trend in combined Atchafalaya-Mississippi River freshwater 

discharge, the amount of freshwater delivered to the Louisiana-Texas continental shelf 

would have increased by 13% (65/57.5 = 1.13). Second and more importantly, it reveals 

that in the 1920s, the Atchafalaya River contributed about one quarter (15/57.5 = 0.26) of 

the freshwater discharge to the Louisiana-Texas continental shelf. Between 1920 and 

about 1960, the Atchafalaya River’s contribution markedly increased to about one half 

(30/65 = 0.46) of the freshwater discharge to the Louisiana-Texas continental shelf. 

While this probably made the Louisiana-Texas continental shelf more prone to hypoxia, 

the timing of this change occurred 15 to 20 years earlier than the onset of regular summer 

hypoxia (Section 2.1.1). 

 Future physical modeling studies are needed to investigate the effects of past and 

proposed future changes in the distribution of freshwater flows, including inputs to 

Atchafalaya Bay some 200 km to the west of the Mississippi River delta, on changes in 

the spatial distribution of surface salinity, temperature, and stratification on the 

Louisiana-Texas continental shelf and on the Mississippi Sound to the east of the birdfoot 

delta. Physical oceanographic models that can adequately answer such questions about 

the impacts of flow diversions already exist but have only been run using the post-1970s 

flow conditions (30% Atchalafaya River, 70% Mississippi River). One such modeling 

study by Hetland and DiMarco (2007) suggests that the freshwater plumes from the 

Atchafalaya and Mississippi Rivers are often distinct from one another (Figure 5) and 

that both contribute significantly to the development of hypoxia (Figure 1) on the shelf 

through their influence on stratification and nutrient delivery (Rabalais et al., 2002a). In 

addition, maps of observed surface salinity and satellite images of chlorophyll (e.g., 

figure 9), show the same result. It thus appears likely that increases in freshwater 

discharge from the Atchafalaya River and resulting increased stratification from the early 

1900’s to the mid-1970’s have increased the area of the Louisiana-Texas continental shelf 

that is prone to bottom layer hypoxia. 
26

Figure 5: Modelled surface salinity showing the freshwater plumes from the Atchafalaya and Mississippi 

Rivers during upwelling favorable winds (top panel) and during downwelling favorable winds 8 days later 

(bottom panel). Adapted from Hetland and DiMarco (2007). 

 Recently evolved plans for protecting coastal Louisiana (CPRA, 2007) propose 

significant diversions of the water, nutrients, and sediment outflow from the Mississippi 

River into the Gulf. Figure 6 illustrates a diversion scenario that involves redirecting a 

large part of the outflow into shallow bays upstream of the present day “bird’s foot” 

delta. This scenario could alter the shelf hydrodynamics, particularly if more of the 

buoyancy is directed into shallow water instead of the deep water off the active river 

mouths, which are near the shelf edge. It is important that three-dimensional numerical 

circulation models be applied to these scenarios. Future management strategies may be 

able to utilize engineered modulations of the timing of freshwater releases to coincide 

more closely with more energetic waves and current conditions, thereby reducing the 

strength of stratification (i.e., Ri). This approach will, of course, rely on engineering 

innovations and effective diversion management. The opportunity exists for EPA and 

other federal and management agencies to urge flow diversion strategies that also 

consider the goal of reducing the volume and bottom area of hypoxic waters on the 

NGOM shelf without endangering other estuarine and coastal waters. The CPRA/U.S. 
27

Army Corp of Engineers proposals also highlight the need for interagency coordination 

and for an integrated approach to management strategies for jointly addressing multiple 

issues including hypoxia, coastal protection, and coastal inundation. 

Figure 6: Proposed diversions of Mississippi effluents for coastal protection. From Coastal Protection and 

Restoration Authority (CPRA) of Louisiana, 2007 Integrated Ecosystem Restoration and Hurricane 

Protection: Louisiana’s Comprehensive Master Plan for a Sustainable Coast. CPRA, Office of the Governor 

(La) 117 pp. 

Zones of hypoxia controls: 

 The resulting stratified region influenced by the Mississippi and Atchafalaya 

River plumes exerts strong control on the extent and spatial distribution of hypoxia and is 

an important factor in determining where hypoxia may occur (Rabalais and Turner, 

2006). The buoyancy fluxes from the rivers also contribute to regional circulation in the 

form of baroclinic flows (Morey et al., 2003a, 2003b). Following a similar line of 

reasoning used in earlier work by Rhoads et al. (1985) off the mouth of the Changjiang 

(Yangtze) River, Rowe and Chapman (2002) defined three zones of hypoxia control in 

the NGOM. The boundaries between these three zones are admittedly fuzzy, and change 

through time; however Figure 7 illustrates the SAB Panel’s view of these concepts as 

represented by 4 zones. In zone 1, which is most proximal to river mouth sources, 

strongly stratified and light- as well as nutrient-limited, respiration of organic carbon 

coming both directly from the river efflux and from nutrient-dominated eutrophication 

dominates. The relative importance of these organic carbon sources as the cause of 

hypoxia remains somewhat uncertain, although the model of Green et al. (2006b) 
28

indicates a major dominance by in situ phytoplankton production even in the immediate 

plume of the Mississippi River. In the intermediate zone 2, stratification is also strong; 

light limitation is less than in zone 1; very high rates of phytoplankton production occur; 

and water column respiration fuels bottom layer hypoxia. Farther along the coast from 

the river mouths but within the low-salinity coastal plume (zone 3), local phytoplankton 

production is less, but labile organic matter may have been imported from zone 2 and 

deposited on the bottom. In zone 3, stratification remains strong, and oxygen 

consumption in the sediment is more important than water column respiration in driving 

hypoxia. Zone 4 depicts the highly productive, coastal current, as suggested by Boesch 

(2003). 

 Boesch (2003) strongly criticized the physical, biological and chemical reasoning 

behind the delineation of the Louisiana-Texas continental shelf into these three distinct 

zones of hypoxia control. He also argued that these zones did not capture well the 

physics and biology of the Louisiana coastal current, which is characterized by low 

salinities and high nutrient and chlorophyll levels (Wiseman et al., 2004). Nevertheless, 

Rowe and Chapman (2002) stimulated new research into the role that stratification plays 

in the reduction of vertical mixing rates and the flux of oxygen through the pycnocline in 

the regions of the Louisiana-Texas continental shelf under the influence of the 

Mississippi and Atchafalaya River plumes. Using realistic three-dimensional physics 

(equation 1) with simple representations of water column and benthic respiration for the 

zones A, B and C of Rowe and Chapman (2002), Hetland and DiMarco (2007) were able 

to represent the bottom area, thickness, and volume of hypoxic waters over the NGOM 

fairly well. 
29

Figure 7: An illustration depicting different zones (Zones 1-4, numbered above) in the NGOM during the 

period when hypoxia can occur. These zones are controlled by differing physical, chemical, and biological 

processes, are variable in size, and move temporally and spatially. Diagram created by D. Gilbert. 

 So far as we are aware, time series measurements of physical oceanographic 

parameters are inadequate to support or refute hypotheses regarding changes in shelf 

circulation, stratification, and vertical mixing during the 20th century. Initial planning for 

a Gulf of Mexico Coastal Ocean Observing System (GCOOS) has begun (for additional 

information see: http://www.gcoos.org). As these GCOOS plans continue to evolve and 

implementation begins over the next few years, it is important that physical parameters 

relevant to oxygen dynamics be included among the measurements. Empirical 

parameterizations of vertical eddy diffusivity Kz as a function of vertical shear and 

density stratification are available for shallow continental shelf environments 

(MacKinnon and Gregg, 2005). These parameterizations enable quantification of vertical 

mixing (term 4 in equation 1) with vertical shear measurements from moored Acoustic 

Doppler Current Profilers (ADCPs) and vertically profiling conductivity, temperature, 

and depth instrumentation (CTDs) tethered on a cable. Ship-based microstructure 

measurements of the turbulent rates of dissipation of velocity, salinity, and temperature 

fluctuations (Gregg, 1999) should also be conducted occasionally to complement the 

moored ADCP and profiling CTD measurements. Physics-based models of ocean mixing 

and turbulence exist today and are part of 3-D circulation models (Mellor and Yamada, 
30

1982). These models need to be rigorously tested using ADCP, CTD, and microstructure 

data because vertical mixing is the most important physical process to model correctly 

when hypoxia is under consideration. 

Shelf circulation: local versus regional 

 Circulation in the NGOM can be considered on two scales: Gulf-wide deep-sea 

circulation and shelf circulation near the coast. Among the most prominent features of 

the large-scale Gulf-wide circulation are the Loop Current and the Loop Current Eddy 

System (Oey et al., 2005a, 2005b). Although these features impinge on and affect the 

outer shelf, Rabalais et al. (1999) conclude that local wind forcing and buoyancy are 

more important to shelf circulation inshore of the 50 meter isobath. Direct ship-board 

observations by Jarosz and Murray (2005) during five separate cruises led those authors 

to conclude that the momentum balance on the inner and mid shelf to the west of the 

active birdfoot delta is indeed dominated by wind stress. During summer, alongshore 

sea-surface slope caused by buoyancy forcing was also important in forcing currents. On 

the 20 m isobath off Terrebonne Bay, ADCP measurements (Wiseman et al., 2004) show 

periods of several days with negligible vertical shear followed by other periods of a few 

days with much more elevated vertical shear and reduced density gradients, suggestive of 

more intense vertical mixing. 

 Several physical oceanographic models taking into account the crucial baroclinic 

effects that typify the Louisiana-Texas continental shelf are now available (e.g., Morey et 

al., 2003a, 2003b; Zavala-Hidalgo et al., 2003). The model results of Hetland and 

DiMarco (2007) show that the plume from the Mississippi River, which enters the shelf 

near the shelf edge, forms a recirculating gyre in Louisiana Bight and does not interact 

with the seabed, whereas the Atchafalaya River plume interacts with the shallow coastal 

topography (Hetland and DiMarco, 2007). Both plumes respond directly to local winds 

and are advected seaward during upwelling-favorable winds (Figure 5). The distinct 

plumes from the Mississippi and Atchafalaya Rivers influence the spatial pattern of 

bottom hypoxia on the Louisiana-Texas continental shelf. This influence is clearly seen 

on the 1985-2005 map of hypoxia frequency of occurrence (Figure 1) and is even more 

obvious in certain years (e.g., 1986, Rabalais and Turner 2006). Given this interaction, 

planned diversions of Mississippi River and Atchafalaya River flow may alter shelf 

circulation and the spatial pattern of bottom hypoxia. Applications of 3-D baroclinic 

models to future scenarios such as that portrayed in Figure 6 are thus important to 

planning for future strategies for coastal restoration (CPRA, 2007). 

 In their analysis of low-frequency (occurring over a time scale greater than 24 

hours) currents over the shelf, Nowlin et al. (2005) distinguished between currents that 

respond within the “weather band” of 2-10 days and those within the mesoscale band of 

10-100 days corresponding to large-scale eddies off the shelf. Inshore of the 50 m 

isobath, the local winds within the weather band dominated and drove currents from east 

to west during non-summer months influenced by the passage of frontal systems. Current 

fluctuations seaward of the 50 m isobath were primarily within the mesoscale band and 

predominantly oriented from west to east but with high variability. Along-shelf and 
31

across-shelf currents in the upper layer over the inner shelf, as reported by Nowlin et al. 

(2005), averaged about 10 cm/s and 1 cm/s, respectively. Over the outer shelf and near 

the seabed, flows were weaker. 

Key Findings and Recommendations 

 The SAB Panel finds that 20th century changes in the hydrologic regime of the 

Mississippi and Atchalafaya Rivers and the timing of freshwater inputs to the LouisianaTexas continental shelf have likely increased the shelf area with potential for hypoxia, 

although these changes occurred mostly from the 1920s to the 1960s, before the 

measured onset of hypoxia in the mid-1970s. Additional work is needed to advance the 

understanding of the relative importance of physical factors in the formation of hypoxia 

in the NGOM. The SAB Panel therefore provides the following recommendations. 

• The development of a new suite of models that integrate physics and 

biogeochemistry should be encouraged and supported. This suite should include 

multiple types of models [i.e., relatively simple models such as those developed 

by Scavia et al. (2003) as well as more complex three-dimensional types such as 

Hetland and DiMarco (2007)]. 

• A comparative impact study of past, present, and future river flow diversions and 

scenarios of altered nutrient supply to the river mouths should be encouraged and 

supported. Three-dimensional hydrodynamic modeling studies are needed to 

compare the spatial distribution of salinity and stratification with 15% (early 

1900’s) and 30% (post-1970’s) Atchafalaya River contributions to the combined 

Atchafalaya-Mississippi River outflow. Coupling of this three-dimensional 

hydrodynamic model with a biogeochemical model would allow quantification of 

the impacts of past river flow diversions on the spatiotemporal extent of hypoxia. 

In addition, to anticipate the possible effects of proposed future effluent diversion 

plans via rerouted deltaic distributaries (CPRA, 2007), these three-dimensional 

biogeochemical and baroclinic shelf circulation models need to be applied to 

scenarios such as that shown in Figure 6 while also considering the effects of 

nutrient-rich Mississippi River waters discharged into local bays and estuaries. 

• Emerging coastal ocean observing and predicting systems in the Gulf of Mexico 

(http://www.gcoos.org) should be encouraged to measure and disseminate 

information needed by hypoxia modelers and those charged with adaptive 

management. Direct measurements of physical and biogeochemical parameters as 

well as direct time series measurement of dissolved oxygen in the bottom 

boundary layer should be routinely provided by the next generation of shelf 

moorings. 

• Studies of turbulent mixing processes involving the effects of stratification over 

the Louisiana-Texas shelf with instruments and techniques capable of quantifying 

turbulent dissipation rates of velocity, salinity, and temperature fluctuations 

should also be encouraged. Studies of the importance of lateral mixing processes 

should be encouraged. 
32

2.1.3. Role of N and P in Controlling Primary Production 

Nitrogen and phosphorus fluxes to the NGOM--background 

 Excessive nutrient loading, dominated by discharge from the MARB, enhances 

planktonic primary production in the shallow near-shore receiving waters of the NGOM 

(Lohrenz et al., 1990, 1992; Turner and Rabalais, 1994; Rabalais et al., 1999a). The 

nutrients of concern are nitrogen (N), phosphorus (P), and silicon (Si) in the form of 

silicate. Appendix B provides general information on the nitrogen, phosphorus and silica 

cycles. Both primary productivity and phytoplankton biomass are stimulated by these 

nutrient sources (Lohrenz et al., 1990, 1992; Ammerman and Sylvan, 2004; Sylvan et al., 

2006). The spatial and temporal extent and magnitudes of this stimulation vary 

significantly, and their patterns and size appear to be related to 1) amounts of freshwater 

discharge and their nutrient loads; 2) the nature and frequencies of discharge (i.e., acute, 

storm- and flood-based versus more gradual, chronic, seasonal discharge); and 3) the 

direction and spatial patterns of discharge plumes as they enter and disperse in the 

NGOM (Justic et al., 1993; Lohrenz et al., 1994; Rabalais et al., 1999b). The Integrated 

Assessment concluded that N loading from the MARB was the primary driver for hypoxia 

in the NGOM. Since the Integrated Assessment, however, considerable knowledge has 

been gained concerning the processes that influence primary production and the relative 

importance of elements other than N as is discussed below. 

 A proportion of the freshwater discharge transits via freshwater and coastal 

wetlands and coastal groundwater aquifers, which modify the concentrations and total 

loads of nutrients entering the NGOM (Day et al., 2003; Turner, 2005). The extent to 

which wetlands alter nutrient loads and the effects wetland losses have had on changes in 

nutrient processing and loading are subjects of considerable debate (Mitsch et al., 2001; 

Day et al., 2003; Turner, 2005). Nutrients can also enter this region from deeper offshore 

sources, by advective transport over the shelf, a modified form of “upwelling” (Chen et 

al., 2000; Cai and Lohrenz et al., 2005), although this input is estimated to be only 7% of 

the nitrogen coming down the Mississippi River (Howarth, 1998). Lastly, nutrients can 

be derived from atmospheric deposition directly onto nutrient-sensitive NGOM waters 

(deposition onto the MARB and subsequent downstream export to the Gulf is considered 

in later sections). For nitrogen, this direct deposition is estimated to be 13% of the 

amount of nitrogen that flows down the river (Howarth 1998). 

 Historic analyses indicate a great deal of variability in seasonal, interannual and 

decadal-scale patterns and amounts of freshwater and nutrient discharge to the NGOM 

(Turner and Rabalais, 1991; Rabalais et al., 2002a). As a result, primary productivity and 

phytoplankton biomass response can vary dramatically on similar time scales, which 

poses a significant challenge to interpreting trends in nutrient-driven eutrophication in the 

NGOM as in other systems (Harding, 1994; Boynton and Kemp, 2000; Paerl et al., 

2006b). Furthermore, in the turbid and highly colored waters (containing colored 

dissolved organic matter or CDOM) of the river plumes entering the NGOM, nutrient and 

light availability strongly interact as controls of primary production and biomass. These 
33

interactive controls modulate the relationships between nutrient inputs and phytoplankton 

growth responses in this region (Justic et al., 2003a, 2003b; Lohrenz et al., 1994). 

Ultimately these interactions affect the formation and fate of autochthonously-produced 

organic carbon that provides an important source of the “fuel” for bottom water hypoxia 

in this region. 

N and P limitation in different shelf zones and linkages between high primary production 

inshore and the hypoxic regions further offshore 

 Physically, chemically and biologically, the NGOM region is highly complex, and 

nutrient limitation reflects this complexity. Along the freshwater to full-salinity 

hydrologic continuum representing the coastal NGOM influenced by river discharge, 

ratios of nutrient concentrations vary significantly, both in time and space. For example, 

depending on the season, specific hydrologic events and conditions (storms, floods, 

droughts), molar ratios of total N to P (N:P) supplied to these waters can vary from over 

300 to less than 5 (Turner et al., 1999; Ammerman and Sylvan, 2004; Sylvan et al., 2006; 

Turner et al., in press). Furthermore, additional environmental factors, such as flushing 

rate (residence time), turbidity and water color (light limitation), internal nutrient 

recycling, and vertical mixing strongly interact to determine which nutrient(s) may be 

controlling primary production (Lohrenz et al., 1999b). Compounding this complexity is 

the frequent spatial separation between high nutrient loads, the zones of maximum 

productivity and hypoxia (e.g., Figure 7). Conceivably, primary production and algal 

biomass accumulation limited by a specific nutrient in the river plume region near shore 

may constitute the “fuel” for hypoxia further offshore in the next zone, where 

productivity in the overlying water column may be limited by another nutrient. 

Limitation by different nutrients in different areas appears to be the case during the spring 

to summer transitional period, when primary production in the river plume region near 

shore is P limited (Lohrenz et al., 1992, 1997; Ammerman and Sylvan, 2004 Sylvan et 

al., 2006), but offshore productivity is largely N limited (Lohrenz 1992, 1997; Dortch and 

Whitledge, 1992). The relevant questions concerning causes of hypoxia are what are the 

relative amounts of inshore river plume (largely P-limited) versus offshore (largely Nlimited) productivity and what roles do these different sources of productivity play in 

“fueling” hypoxia. 

 Early work on NGOM nutrient limitation tended to focus on the waters overlying 

the hypoxic zone; typically, these waters are over the shelf but farther offshore than the 

river plume waters. Stoichiometric N:P ratios indicated that, during summer months 

when hypoxia was most pronounced, N should be the most limiting nutrient (Justić et al., 

1995; Rabalais et al., 2002a). This work has been the basis for the general conclusion 

that N is most limiting, and that reductions in N loading would be most effective in 

reducing “new” carbon (C) fixation and resultant phytoplankton biomass supporting 

hypoxia (Rabalais et al., 2002a, 2004). This conclusion, coupled with the nutrient 

loading trend data over the past 40-50 years, which showed N loading increasing more 

rapidly than P loading, has formed the basis for arguing that N input reductions would be 

most effective in reducing the eutrophication potential and hence formation of “new” C 

supporting hypoxic conditions. The 2000 report from the National Academy of Sciences’ 
34

Committee on Causes and Management of Coastal Eutrophication (National Research 

Council, 2000) concluded that nitrogen is the primary cause of eutrophication in most 

coastal marine systems in the U.S. at salinities greater than 5 – 10 parts per thousand 

(ppt), including the NGOM. 

 While it is likely that N limitation characterizes coastal shelf and offshore waters, 

more recent nutrient addition bioassays (Ammerman and Sylvan, 2004; Sylvan et al., 

2006) and examinations of nutrient stoichiometric ratios have shown that river plumeinfluenced inshore productivity appears to be more P limited, especially during periods of 

highest productivity and phytoplankton biomass formation (Feb-May) (Figure 8) when 

freshwater discharge and total nutrient loading are also highest (Lohrenz et al., 1999a, 

1999b; Sylvan et al., 2006). 

Figure 8: Response of natural phytoplankton assemblages from coastal NGOM stations to nutrient 

additions, March through September. All experiments, except those done in September, indicate a strong 

response to P additions. Taken from Sylvan et al., 2006. 

 The strong P limitation during this period appears to be a result of the very high 

rates of N loading that have increased more rapidly than P loading over recent history 

(the past 50 years) (Turner and Rabalais, 1991; Turner et al., 1999). This situation is 

exacerbated during periods of high freshwater runoff, which typically contain very high 

N:P ratios. Primary productivity in the river plume region near shore tends to shift into a 

more N limited mode once freshwater discharge decreases during the drier summer-fall 

period (June-October). However, total primary production and phytoplankton biomass 
35

accumulation are far lower during this more N-limited period than during the earlier Plimited period. Overall, maximum “new” organic C formation in recent years tends to 

coincide with periods of highest N:P, which are P limited (Lohrenz et al., 1992, 1997, 

1999a; Ammerman and Sylvan, 2004; Sylvan et al., 2006). 

 Field data and remote sensing imagery indicate that in situ phytoplankton biomass 

(as chlorophyll a) concentrations can be quite high in river plume-influenced inshore 

waters that have been shown to be P limited. This pattern is evident in Figure 9, an 

image provided by the National Oceanic and Atmospheric Administration Sea-viewing 

Wide Field-of-view Sensor Project (NASA-SeaWiFS, 2007). Therefore, the following 

question emerges. What is the spatiotemporal linkage of this P-limited high primary 

production and phytoplankton biomass accumulation to hypoxic bottom waters located 

further offshore? Furthermore, what are the relationships between N-limited production 

later in the summer and hypoxic conditions, which typically are most extensive during 

this period? These potential “relationships” are complicated by the fact that there are 

strong, co-occurring physical drivers of hypoxia, including vertical density stratification 

and respiration rates, which tend to be maximal during periods of maximum development 

of hypoxia (c.f. Rowe and Chapman, 2002;Wiseman et al., 2004; Hetland and DiMarco, 

2007; DiMarco et al., submitted). 

Figure 9: NASA-SeaWiFS image of the Northern Gulf of Mexico 

recorded in April, 2000. This image shows the distributions and 

relative concentrations of chlorophyll a, an indicator of phytoplankton 

biomass in this region. Note the very high concentrations (orange to 

red) present in the inshore regions of the mouths of the Mississippi and 

Atchafalaya Rivers. 
36

 There are likely to be periods when both P and N are supplied at very low levels 

and co-limit phytoplankton production. These periods occur during the transition from 

spring to summer. A similar condition is observed in large estuarine systems with a 

history of eutrophication, such as Chesapeake Bay (Fisher et al., 1992). Spatially, the 

upstream, freshwater segments of Chesapeake Bay tend to be most P limited, especially 

during spring runoff conditions, while the more saline down-estuarine waters tend to be 

most N limited. In Chesapeake Bay, the more turbid upstream freshwater component 

tends to exhibit interactive light and P limitation or N+P co-limitation (Fisher et al., 

1992; Harding et al., 2002). Farther downstream, light limitation plays a less important 

role. This scenario could prove similar to the riverine-coastal continuum in the NGOM, 

where the most turbid upstream river plume waters are likely to exhibit the highest 

probability for light-nutrient interactive limitation of primary production (Lohrenz et al., 

1999a, b). 

 While bioassay data tend to indicate P limitation during springtime in the lower 

salinity portions of this continuum and N and P co-limitation and N limitation in the more 

saline offshore waters during summer months, the bioassays do not account for sedimentwater column exchange because sediments are excluded during the course of incubation. 

It is possible, although unlikely because of short incubation times, that sediment-water 

column P cycling in the shallow NGOM water column may minimize P limitation in situ. 

In order for this scenario to be operative, parallel N recycling would have to be far less 

efficient than P cycling, which numerous studies suggest is the case (Gardner et al., 1994; 

Bode and Dortch, 1996; Pakulski et al., 2000; Wawrik et al., 2004; Jochem et al., 2004; 

Cai and Lohrenz, 2005). Bioassay-based N limitation results might also be influenced by 

the elimination of “internal” sediment-water column N recycling, although this situation 

seems unlikely as well, especially if denitrification is operative (Childs et al., 2002). 

Sediment-based denitrification would lead to N “losses” from the system, thereby 

exacerbating N limitation. This influence would not be captured in bioassays, which 

isolate the sediments from the water column during incubation. The relatively short 

incubation times of bioassays probably preclude these potential artifacts. They offer a 

“snapshot” of nutrient limitation to complement longer-term, ecosystem-scale 

assessments. 

 The degree of N and P limitation can be calculated from bioassays, and the data 

can be used to create ratios of N and P limitation (Dodds et al., 2004). Interestingly, N 

and P limitation inferred from stoichiometric ratios of soluble (and hence biologicallyavailable) inorganic or total N or P concentrations and inputs (loads) tends to confirm 

bioassay-based conclusions concerning specific nutrient limitations. For example, 

inshore, river-influenced waters exhibit quite high molar N:P ratios, often exceeding 50 

[Nutrient Enhanced Coastal Ocean Productivity (NECOP) Reports, NOAA, 2007]. 

Nutrient addition bioassays initially conducted in these waters by Lohrenz et al. (1999a) 

and more recently by Sylvan et al. (2006), consistently revealed P limitation, especially 

during spring periods of maximum primary production and phytoplankton biomass 

accumulation. These same studies also indicated a tendency towards N and P colimitation and exclusive N limitation during later summer months, when soluble and total 

N:P values dipped below 15. It should also be noted however that rates of primary 
37

production and phytoplankton biomass during this more N-limited period are at least 

five-fold lower than spring values, according the Gulf of Mexico NECOP data (Lohrenz 

et al., 1999a, 1999b). Sylvan et al. (2006) point out that P-limited spring production of 

“new” C may play a proportionately greater role than N-limited summer production as a 

source of “fuel” supporting hypoxia in the NGOM. The degree and extent to which C 

from this nutrient-enhanced elevated spring production is transported and accounts for 

summer hypoxia need to be quantified. Developing an understanding of processes that 

link zones and periods of high primary production and phytoplankton biomass to zones 

exhibiting bottom water hypoxia is a fundamentally important and challenging area of 

research. Such research is necessary to improve understanding of the linkage between 

nutrient-enhanced production and bottom water hypoxia in the NGOM. Extrapolation of 

C production to hypoxia data along the entire riverine-coastal shelf continuum, where 

zones and periods of maximum productivity and bottom water hypoxia do not necessarily 

coincide or overlap, depends on knowing C transport and storage (including burial), 

internal nutrient, and C cycling and C consumption (heterotrophic metabolism and 

respiration) processes along this continuum (Redalje et al., 1992; Cai and Lohrenz, 2005). 

Quantifying the links between locations and periods of specific nutrient limitation (or 

stimulation) of production and the fate of this production relative to hypoxia will 

contribute to long-term, effective nutrient management strategies for this region. 

Key Findings and Recommendations

 The SAB Panel finds that there is compelling evidence that the near shore 

Mississippi/Atchafalaya River plume-influenced waters are P limited and P-N co-limited 

during the spring periods of highest primary production. Nitrogen limitation of primary 

production prevails during summer periods. Recent research results indicate that the 

spring period of maximum primary production is P-limited in at least the plumes of the 

rivers, largely due to excessive N input. As a result of this man-made imbalance in 

nutrient loading during this crucial period, P availability plays an important role in 

contributing to the production of “new” organic carbon in the spring time and quite likely 

contributing in a major way to the “fueling” of summer hypoxia in the NGOM. 

However, as stressed elsewhere in this report, there is great uncertainty over the coupling 

in space or time of phytoplankton production and its decomposition leading to hypoxia. 

Therefore, a better understanding of the spatial extent and temporal patterns of these 

nutrient limitations is needed. The SAB Panel recommends that the following work be 

undertaken to advance knowledge of the importance of nutrient limitation and colimitation as factors in the formation of Gulf hypoxia. 

• Research should be conducted to develop a more complete understanding of the 

spatial and temporal linkages between river plume-influenced inshore P (in 

spring) and/or N limited (in summer) primary production, and offshore coastal 

shelf, more N-limited production, as well athe fate of C produced in each zone 

throughout the year. 

• Research should be conducted to link near inshore river plume-influenced 
38

production in time and space to O2 depletion farther offshore. Green et al. 

(2006b) suggest that the small region that the central Mississippi River plume 

could supply is responsible for about 25% of the C necessary to fuel hypoxia. 

The role of the Atchafalaya plume and other riverine influenced, inshore high 

productivity regions in offshore hypoxia needs to be clarified. 

• Research should be conducted to address the following questions. How closely 

linked are the periods of high productivity and hypoxic events throughout the 

regions in which they occur? What is the lag between C production and its 

ultimate degradation? 

2.1.4. Other Limiting Factors and the Role of Si 

 While excessive N and P loading are implicated in eutrophication of the NGOM, 

these nutrients also play a role in the balance, availability and ecological manifestations 

of other potentially-limiting nutrients, most notably Si. In the Mississippi River plume 

region, N is supplied in excess of the stoichiometric nutrient ratios needed to support 

phytoplankton and higher plant growth (i.e., Redfield ratio, Redfield, 1958). If N overenrichment persists for days to weeks, other nutrient limitations may, at times, result and 

seasonally dominate; the most obvious and important is P limitation, which has recently 

been demonstrated in bioassays (Ammerman and Sylvan, 2004; Sylvan et al., 2006). In 

addition to P limitation, N and P co-limitation and Si limitation (of diatom growth) have 

been observed in the fresh and brackish water components of riverine plumes that can 

extend more than 100 km into the receiving waters (Dortch and Whitledge, 1992; 

Lohrenz et al., 1999a; Dortch et al., 2001). A similar scenario is evident in the 

Chesapeake Bay, where elevated N loading accompanying the spring maximal freshwater 

runoff period increases the potential for P limitation (Fisher and Gustafson, 2004). The 

biogeochemical and trophic ramifications of such shifts are discussed below. 

Can increased N:Si and P:Si fuel an increased microbial loop and exacerbate hypoxia? 

 With regard to nutrient primary production interactions, it is important to know 

who the dominant primary producers are, where they reside, what their contributions to 

new production are, and what their fate is. In NGOM waters downstream of the rivers, 

wetlands and intertidal regions, microalgae are by far the dominant primary producers 

(Lohrenz et al., 1992, 1997; Redalje et al., 1992; Rabalais et al., 1999a). The microalgal 

communities are dominated by phytoplankton (Redalje et al., 1994a, 1994b; Chen et al., 

2000) although benthic microalgal communities can also be important sites of primary 

production and nutrient cycling, especially in near-shore regions (Jochem et al., 2004). 

As nutrient loads and limitations change over time and space, the proportions of 

planktonic versus benthic microalgae may also change; i.e., as nutrient inputs are reduced 

and planktonic primary production is reduced, the microalgal community may shift to a 

more benthic dominated one. This process could yield significant implications for 
39

biogeochemical (nutrients, carbon and oxygen) cycling and trophodynamics (Rizzo et al., 

1992; Darrow et al., 2003). 

 Historic and contemporary evidence supports the contention that 

anthropogenically and climatically-induced changes in N and P loading have increased 

NGOM primary productivity and phytoplankton biomass and altered phytoplankton 

community composition. There are several reasons why phytoplankton community 

composition may have been altered by changes in nutrient loading: 1) competitive 

interactions among phytoplankton taxa based on varying nutrient supply rates and 

differing affinities for nutrient uptake and assimilation (i.e., varying nutrient uptake 

affinities and kinetics); 2) competitive interactions based on the relationships between 

nutrient supply rates and photosynthetically available light (i.e., low versus high light 

adapted taxa); 3) competitive interactions based on changes in N versus P supply rates 

(e.g., differential N versus P uptake capabilities and selection for nitrogen fixing 

cyanobacteria); 4) competition based on the ratios of N and P versus Si (silicious versus 

non-silicious taxa and heavily- versus lightly-silicified diatoms); 5) differential grazing 

on phytoplankton taxa (top-down controls); and 6) nutrient-salinity controls (interactive 

effects of changes in freshwater discharge on NGOM salinity and nutrient regimes due to 

climatic and watershed hydrologic control changes). Each set of controls can influence 

the amounts and composition of primary producers. These controls can also interact in 

time and space, greatly compounding and confounding the interpretation of their 

combined effects. 

 One important aspect of differential nutrient loading is the well-documented 

increase in N and P relative to Si loading. While N and P loads tend to reflect human 

activities in and alterations of the watershed, Si loads tend to reflect the mineral (bedrock 

and soil) composition of the watershed; a geochemical aspect that is less influenced by 

human watershed perturbations. Agricultural, urban and industrial development and 

hydrologic alterations in the MARB have led to dramatic increases in N and P relative to 

Si loading. In addition, the construction of reservoirs on tributaries of these river systems 

has further exacerbated this situation by trapping Si relative to N and P. This 

anthropogenic biogeochemical change has been shown to alter phytoplankton community 

structure (i.e., away from diatom dominance), with subsequent impacts on nutrient and 

carbon cycling and food web dynamics (Humborg et al., 2000; Ragueneau et al., 2006a, 

2006b). The overall result has been an increase in N:Si and P:Si ratios that can influence 

both the amounts and composition of phytoplankton; including potential shifts from 

diatoms to flagellates and dinoflagellates (Turner et al., 1998; Rabalais and Turner, 2001; 

Justic et al., 1995). Diatoms are a highly desired food item for a variety of planktonic 

and benthic grazers, including key zooplankton species serving an intermediate role in 

the NGOM food web (Dagg, 1995). The dinoflagellates, cyanobacteria and even a few 

diatom species, while serving important roles in the food web, also contain species that 

may be toxic and/or inedible (Anderson and Garrison, 1997; Paerl and Fulton, 2006). 

Some of these species can rapidly proliferate or “bloom” under nutrient sufficient and 

enriched conditions, and thus constitute harmful algal bloom (HAB) species. Toxicity 

may directly and negatively impact consumers of phytoplankton as well as higher-ranked 

consumers, including finfish, shellfish and mammals (including humans). If non-toxic 
40

but inedible (due to size, shape, coloniality) phytoplankton taxa increase in dominance, 

trophic transfer may be impaired. Planktonic invertebrates, shellfish, and finfish 

consumers (whose diets are highly dependent on the composition and abundance of 

specific phytoplankton food species and groups) may then be affected (Turner et al., 

1998). This could have consequences for C flux, with a relatively higher fraction of C 

being processed through microbial pathways (i.e., the “microbial loop”) or sedimented to 

the bottom. In either case, a greater fraction of the primary production would remain in 

the system, as opposed to being exported out of the system by transfer to higher trophic 

level and fisheries. The net result would be more C metabolized within the system, 

leading to enhanced oxygen consumption and increased hypoxia potentials. 

Key Findings and Recommendations 

 Research has shown the potential importance of silicate in structuring 

phytoplankton communities. Based on this finding, the SAB Panel offers the following 

recommendation. 

• The potential for silicate limitation and its effects on phytoplankton production 

and composition on the Louisiana-Texas continental shelf should be explored 

when carrying out experiments on the importance of N and P as limiting factors 

and when considering nutrient management scenarios. 

2.1.5. Sources of Organic Matter to the Hypoxic Zone 

 As noted earlier, the physical and geomorphological conditions found along the 

Louisiana coast make the NGOM prone to hypoxic conditions if there is an organic 

matter supply sufficient to consume deep water dissolved oxygen (DO) at rates exceeding 

DO replenishment rates. Ecosystems such as the NGOM shelf have available to them an 

array of organic matter sources, including those transported from the basin by rivers and 

those produced in-situ. These include particulate and dissolved organic carbon/colored 

dissolved organic matter (POC and DOC/CDOM) from terrestrial sources in the basin, 

POC and DOC from coastal wetland losses, and in-situ production by phytoplankton, 

macrophytes, and benthic microalgae. 

 The Integrated Assessment largely supported the argument that hypoxia in the 

NGOM was driven by increased N loading to the Gulf of Mexico, which, in turn, 

stimulated increased in-situ phytoplanktonic production of labile (i.e., readily 

decomposed) organic matter. A portion of this organic matter sinks to deeper, subpycnoclinal waters and is used by the heterotrophic community at rates sufficient to 

deplete DO concentrations to hypoxic levels. Emphasis at that time focused on N but 

more recent work has indicated that P also plays a role in regulating organic matter (OM) 

supply from phytoplankton (see Section 2.1.3). In addition, a number of investigators 

have noted that changes in the relative supply rates of N, P and Si lead to changes in 
41

species composition of phytoplankton communities, and this would likely modify some 

aspects of deposition of OM to deep waters. Substantial rates of primary production have 

been measured along the NGOM shelf, and these rates are comparable to those observed 

in other eutrophic coastal systems (e.g., Lohrenz et al., 1990; Lohrenz et al., 1997; Nixon, 

1992). 

 In Rabalais et al. (1999a) and the Integrated Assessment, organic matter from the 

major rivers was discounted as a major source because 1) there have not been changes in 

river OM loads since the beginning of the hypoxic period that account for the current 

hypoxic zone size and expansion; 2) dissolved organic matter (DOM) sources from 

rivers, while large, would need to be converted into particulate forms, with attendant 

losses from this microbial transformation, and hence would be much reduced; 3) much, 

but not all, of this terrestrially derived material is far less labile than phytoplanktonic 

debris and hence is not readily respired at time scales associated with shelf hypoxia 

(weeks to months). Using an estimated annual load of river OM (~ 2.6 x 1012 g C/ yr) 

delivered to an average hypoxic area (15,000 km2

), and assuming that even as much as 

30% of this material were labile, suggests a small impact on DO conditions (~0.3 g O2/m2 

/day). Additionally, while there is substantial POC and DOC coming down the 

Mississippi River, there was undoubtedly far more 100 - 130 years ago when the 

Mississippi River basin was first cleared for agriculture and before the dams in the basin 

were built. While this process apparently has not been modeled in the Mississippi River 

basin, modeling in other basins strongly suggests a huge increase in organic carbon fluxes 

at the time of land-use conversion to agriculture, followed by decreasing fluxes as 

agricultural practices improve (Swaney et al., 1996), and globally the flux of carbon in 

rivers is tied to agricultural land use (Schlesinger and Melack, 1981). This historical 

land-use change may well have contributed to the paucity of low oxygen conditions seen 

in the paleoecological record in the late 1800s (Osterman et al., 2005). Given this 

historical pattern, Mississippi River derived OM is unlikely to be the trigger for the level 

of hypoxia that developed in the NGOM during the past 35 years. This period does 

coincide well with the time N loads increased, due mainly to the use of synthetic N 

fertilizer in the Mississippi River basin. Given experience in many other coastal and 

estuarine regions (e.g., National Research Council, 2000), there are strong reasons to 

believe that in situ NGOM primary productivity exploded in response to increased N 

inputs over this time scale. 

 The influence of organic matter losses from coastal wetlands on coastal hypoxia is 

still debated but seems unlikely to be a primary factor. Whether or not wetlands lose 

more organic C as they degrade is not well known, but at present this also seems unlikely. 

While the timing of wetland loss does not coincide with the onset of hypoxia in the 1970s 

(marsh loss has been occurring since the 1940s), stable isotope and lignin analyses of OM 

over much of the shelf indicates that terrestrially-derived OM is dispersed along and 

across the shelf (Goni et al., 1998; Gordon et al., 2001). However, marsh particulate 

organic material is refractory (i.e., resistant to decay) and does not contribute much to 

hypoxia creation on time scales of weeks to months. Thus, while the conclusion that the 

main OM source fueling hypoxia is in-situ production of marine phytoplankton and that 

this production increased in response to enhanced nutrient loads from the MARB remains 
42

sound, a better understanding of the possible role of other sources would further refine 

understanding of hypoxia. 

Sources of organic matter to NGOM: post 2000 Integrated Assessment 

 Since the Integrated Assessment, there has been substantial research activity in the 

NGOM regarding organic matter sources, characterization of organic matter, and related 

issues. Some of this new work has utilized advanced analytical methods and improved 

field techniques. However, as with the advent of sophisticated imaging devices in 

medicine, where small and interesting structures in the human body can now be readily 

observed but not necessarily interpreted in terms of health threats, in marine waters we 

now have an emerging and more detailed description of the complex mix of organic 

compounds, which has in the past simply been called organic matter. But it is not yet 

clear how important some of this material is with respect to hypoxia issues. This 

elaboration of understanding of OM adds interesting and useful dimensions to this story 

but does not change the basic theme, which is that enhanced phytoplanktonic production, 

based on much increased nutrient loading, is the main biological trigger of NGOM 

hypoxia. 

 In addition, there have been at least two varieties of what can be called synthesis 

studies. Studies of the first variety tend to be “review-like” wherein the growing timeseries of observations and new data have been revisited and/or re-analyzed. Several other 

efforts of this type have also developed revised conceptual models of the role of OM in 

hypoxia, and these will prove especially useful in time. Studies of the second variety, 

and these are rarer, involve development of quantitative budgets or models of various 

sorts. These efforts indicate that the information base regarding many aspects of OM and 

hypoxia is rich enough to begin these more rigorous examinations. But, in virtually all 

these efforts, authors conclude that results are preliminary and that more process-based 

information is critically needed. 

Advances in organic matter understanding: characterization and processes 

 A detailed review of these diverse studies is beyond the scope of this effort. 

However, Table 1 summarizes a selection of those works to provide an indication of the 

diversity of information that is becoming available. Some findings of particular 

relevance to OM sources are provided below: 

• POC associated with sand transport in bottom waters in the lower Mississippi 

River is similar in magnitude to loading of suspended POC (Bianchi et al., 2007). 

• The vertical flux of terrestrially-derived particles in the Mississippi River plume 

was typically very high and mainly deposits locally (Corbett et al., 2004). 

• Recent analyses suggested that woody angiosperm material (13C-depleted) 

preferentially settled within the lower Mississippi River and in the river plume 

(Bianchi et al., 2002). Other work has demonstrated that erosion of relict peat in 
43

transgressional facies of the lower Mississippi River provide a source of “old” 

vascular plant detritus to the river plume (Galler et al., 2003). 

• High sedimentation rates in the river plume result in the formation of mobile mud, 

commonly observed in other large river-ocean interfaces (McKee et al., 2004). It 

is estimated that about 50% of the sediments (and associated OM) delivered to 

this region are temporarily stored near the delta – with a large fraction transported 

along/across the shelf in the benthic boundary layer (Corbett et al., 2004, 2006). 

• Diatom signals in surface sediments suggested possible inputs of riverine diatom 

phytodetritus to the inner shelf (Wysocki et al., 2006). Previous work showed 

higher phytoplankton biomass, mostly as diatoms, than expected in the lower 

river (Dagg et al., in press; Duan and Bianchi, 2006) with conversion, via lysis, to 

DOC. Hence, river nutrients were converted to river phytoplankton biomass and 

then ultimately to DOC, providing a labile food resource for bacterioplankton. 

• An analysis of OM production to the west of the plume found phytoplankton at 

the outer edge of this region declined due to nutrient limitation, microzooplankton 

followed trends in phytoplankton, most particle sinking was associated with 

mesoplankton fecal pellets, phytoplankton-derived DOM reached a peak and was 

correlated with bacterioplankton, and water column recycling was most intense in 

this region (Dagg and Breed, 2003). 

• Estimates suggested 10% to 52% of the DOM in the region west of the plume is 

quite labile (Benner and Opsahl, 2001). More recent data indicated that most 

riverine DOC was photochemically converted to dissolved inorganic carbon 

(DIC) over a period of weeks in this region (Dagg et al., in press). More 

terrestrially-derived components such as lignin had similar fates (Hernes and 

Benner, 2003). 

• Some labile sedimentary organic matter, from in situ diatom production, was 

rapidly (day to weeks) shunted to the Mississippi River Canyon (Bianchi et al., 

2006), essentially bypassing the hypoxic zone to the west. The supply rate of this 

phytodetritus was sufficient to support macrobenthic polychaete populations that 

do not exist in nearshore waters off the Louisiana coast. The removal of labile 

OM by winter season and hurricane events may act as a cleansing mechanism, 

reducing the potential for hypoxia (Bianchi et al., 2006). 

• There are plumes from rivers and local estuaries along the coast containing 

colored dissolved organic matter (Chen and Gardner, 2004). DOC concentrations 

are also generally high (Engelhaupt and Bianchi, 2001) but higher still in the 

Atchafalaya River than the Mississippi River (Chen and Gardner, 2004; Pakulski 

et al., 2000; Bianchi et al., 2004). 

These brief comments hardly do justice to the vast amount of work completed 

since the Integrated Assessment. However, they do provide evidence of improved 

understanding and elaboration of the role of different forms of OM in the NGOM 

ecosystem. 
44

Synthesis efforts regarding organic matter sources 

 In most environmental analyses, synthesis of diverse data sets is essential for 

clarifying cause-effect couplings and sorting out primary from secondary effects. 

Hypoxia and the role of various OM sources in NGOM hypoxia are no exception. 

Fortunately, a variety of descriptive and more quantitative syntheses/reviews have been 

developed since the Integrated Assessment. 

 Several studies, including those of Rabalais et al. (2002), Turner et al. (in press), 

Justic et al. (in press) and Rabalais et al. (2007), largely reaffirm the primacy of river 

nutrients in supporting high rates of in-situ primary production as the dominant source of 

OM supporting intense ecosystem respiration and development of hypoxic conditions. 

Walker and Rabalais (2006) analyzed SeaWiFS algal biomass data in relationship to river 

flow, nitrate loads from rivers and hypoxia. Results confirmed strong relationships 

between nutrient loading and algal biomass distributions; direct relationships to hypoxic 

waters remained elusive for a variety of reasons. The importance of this work lies in the 

fact that the whole hypoxic-prone zone was assessed in a synoptic fashion and data were 

available for both low and high nutrient load periods. Dagg et al. (in press) also reviewed 

data to determine Mississippi River plume contributions to hypoxia. Results were largely 

consistent with those noted above, but Dagg et al. (in press) focused on the important role 

of the plume in both producing and consuming organic matter and dissolved oxygen and 

in building a case for the importance of coastal wetlands as an important organic matter 

source. However, there are problems with the magnitude of wetland OM contributions 

suggested by these calculations, including conversion of wetland sediment losses to OM 

mass, no consideration for on-marsh respiration of this material, and no consideration of 

the refractory nature of the particulate material, a major portion of this OM. Based on 

present understanding of the issue, it seems unlikely that wetland loss could be a prime 

source of OM to the hypoxic zone. 

 Finally, there have been several quantitative assessments of OM for portions of 

the hypoxic zone, and these are emphasized here because it seems that these types of 

syntheses are especially useful in understanding hypoxia and could serve as templates for 

designing future data acquisition programs. Several other studies, including those of 

Rowe and Chapman (2002) and Dagg and Breed (2003) have proposed broader 

conceptual models for the plume and the full hypoxic zone, respectively, and these might 

also be useful in study design and improving our vocabulary when discussing the hypoxic 

zone and the role of various OM sources. Gordon et al. (2001) used a variety of 

measurements to evaluate the distribution and accumulation of organic matter on the 

shelf west of the Atchafalaya River. They reported inputs from rivers and in-situ 

production (in-situ production dominated), estimated OM losses due to water column and 

sediment respiration (OM substrates being marine and riverine, respectively) and longterm burial (< 5% of total inputs). Green et al. (2006b) used careful delineation of the 

Mississippi River turbidity plume coupled to a biological model to investigate OM 

budgets for this zone. They reported that labile OM was mainly from autochthonous 

phytoplankton production and that riverine OM inputs to the plume were three times as 

large but quite refractory. Losses of OM were mainly from microbial respiration, and, 
45

importantly, the plume as a whole was net autotrophic, again suggesting the primacy of 

in-situ production. Finally, while the plume is a small fraction of the full hypoxic zone, 

Green et al. (2006b) estimated that plume derived OM was equivalent to about 23% of 

the OM needed to create observed hypoxia on the full shelf. 

Table 1: A partial summary of papers published following the Integrated Assessment related to sources of organic 

matter to the Gulf of Mexico.

 

General Topics and issues Comments regarding OM/hypoxia Reference 

Landside Sources

POC in river sands Similar in magnitude to suspended POC load in river Bianchi et al., 2002 

Sedimentation of river POC high deposition of terrestrial POC in plume region Corbett et al., 2004 

Relict peats source of old organic matter to plume area Galler et al., 2003 

Seasonal transport of POC fluid muds are transported seasonally to GOM McKee et al., 2004 

Sediment storage and transport seasonal transport of mobile muds from delta to shelf Corbett et al., 2006 

River OM loads DOC and DON loads to GOM Bianchi et al., 2004; Duan et al., 2007 

River inputs transport of river diatoms to plume area Wysocki et al., 2006; Duan and 

Bianchi, 2006 

Terrestrial OM fate of lignin Hemes and Berner, 2003 

Riverine DON photoammonification of DON to DIN Pakulski et al., 2000 

Riverine OM and nutrients Effects of flow through coastal wetlands Xu, 2006 

Riverine DOM CDOM analysis Chen and Gardner, 2004 

Marsh/estuary DOC high DOC concentrations in these systems Engelhaupt and Bianchi, 2001 

OM distribution sources and fate of OM from rivers to shelf Gordon et al., 2001 

Water Column/ Sediment Processes

Flocculation and sedimentation enhanced process in plume area; high rates Dagg et al., 2004 

Light field light absorption/scattering limiting production D’Sa and Miller, 2003 

Plankton characteristics satellite-based relations between N-loads and 

chlorophyll 

Walker and Rabalais, 2006 

Plume budget CO2 budget in plume Cai, 2003 

OM source high rates of plankton production west of plume Dagg et al., in press 

Deposition Influence of larvaceans on deposition Dagg and Brown, 2005 

DOM characteristics Lability of DOM in Region II Benner and Opsahl, 2001 

Sediment DOC Release of DOC from shelf sediments Sutula et al., 2004 

Fate of benthic diatoms Benthic diatom shunted to MR canyon; cleansing effect Bianchi et al., 2006 

Hurricane effects storm transport of deposited materials-decadal scale Corbett et al., 2006 

Sediment processes ammonium flux from sediments important for plankton Eldridge and Morse, 2007 

Plankton composition Diatom occurrence in western regions of hypoxic zone Warwick and Paul, 2004 

Plankton composition microbial processes in shelf waters Liu et al., 2004 

Synthesis/Overviews

OM budget Carbon budget for plume area Green et al., 2006b 

Conceptual model/synthesis planktonic dynamics of region outside plume Dagg and Breed, 2003 

Model analysis differences between water and sediment respiration Hetland and DiMarco, 2007 

Statistical model relates N-load to hypoxia; phytoplankton OM implied Scavia et al., 2003 
46

 

General Topics and issues Comments regarding OM/hypoxia Reference 

Water column synthesis plume contributions to hypoxia; gaps in understanding Dagg et al., in press 

Review/synthesis new monitoring data strengthens nutrient/hypoxia model Rabalais et al., 2007 

Nutrient/Organic loads confirms Integrated Assessment, wetland loss small OM Turner et al., in press 

Forecasting hypoxia examines models and suggests nutrients major driver Justic et al., in press 

Primary production-nitrate model model indicates buffered response to N-load reductions Green et al., in press 

Concepts of hypoxic zones suggests spatial dimensions/processes in hypoxic zones Rowe and Chapman, 2002 

* Entries are shown for a variety of topics and comments are focused on issues related to organic matter in 

the GOM. This table is not a complete summary of all papers published on this subject; rather it provides 

an indication of the great diversity of studies conducted since the Integrated Assessment. 

Key Findings and Recommendations 

 The SAB Panel concludes this section with several findings. First, there is general 

and strong support for the conclusion that riverine nutrients support levels of plankton 

production capable of creating observed hypoxic conditions. However, some aspects of 

the relationship between in-situ phytoplankton production and hypoxia remain uncertain. 

There is need for additional study of the hypoxia issue that emphasizes process studies 

and better coupling of physics to the chemical and biological features of the hypoxic 

zone. The SAB Panel therefore provides the following recommendations. 

• Continued research should be conducted to further elucidate the role of N and P 

from the MARB in stimulating phytoplankton production, the primary drivers 

creating excess OM and thus hypoxia in the Gulf. 

• A series of consistent, well-placed, and well-timed process studies should be 

conducted in the NGOM. Virtually all the OM review/synthesis papers referenced 

above state that their analyses suffer from a lack of pertinent process data. 

• DOM and POM delivered to the NGOM by rivers and from coastal wetland losses 

represent potential OM sources. The weight of evidence currently available 

suggests that it is unlikely these were triggers for hypoxia development or primary 

OM sources for hypoxia maintenance. However, the magnitude of river OM 

sources is large, and hence further characterization of this material is warranted. 

2.1.6. Denitrification, P Burial, and Nutrient Recycling 

 The availability of N and P in an ecosystem is controlled both by external 

loadings and internal biogeochemical processes. Ideally information is needed on the 

load of biologically available nutrients, which is not necessarily well reflected by either 

the load of dissolved inorganic nutrients or the load of total nutrients. Internal 

biogeochemical processes are poorly known for the NGOM. Some, but not all, of the 

dissolved organic nutrients and particle-bound nutrients delivered to coastal waters 
47

become biologically available on ecologically meaningful time scales (days to months). 

In the Mississippi River, the fate of the particle-bound P is of particular interest since it is 

the most common form of P in the river (Sutula et al., 2004). The bioavailability of this 

form of P is low within the freshwater portions of the Mississippi River, but, as the 

particles encounter the increasingly more saline waters of the Gulf of Mexico, the high 

ion abundances of seawater cause much of the adsorbed inorganic P to desorb, converting 

it into highly bioavailable dissolved inorganic P (Fox et al., 1985; Froelich, 1988; 

Howarth et al., 1995; Sutula et al., 2004). In addition, sediment diagenetic processes 

further increase the biological availability of particle-bound P delivered to the Gulf 

(Sutula et al., 2004). 

 For many coastal marine systems, the tendency is for benthic processes to make N 

limitation more prevalent since the N sink through denitrification is relatively larger than 

is the loss of P through permanent sediment burial (National Research Council, 2000; 

Blomquist et al., 2004; Howarth and Marino, 2006). Phosphorus release from sediments 

is frequently less than the rate of P remineralization, due to P adsorption and storage in 

surface sediments (National Research Council, 2000; Howarth and Marino, 2006). 

Variations in P release are probably due to differences in the amount and forms of iron in 

the sediments, the extent of sulfate reduction, and mixing by the benthic fauna, 

particularly as this affects micro-scale variation in pH (Howarth et al., 1995). The 

dynamics of P-sediment exchanges in the Louisiana shelf region are sufficiently complex 

that in a recently published model of sediment diagenesis (Morse and Eldridge, in press), 

P processes were deliberately not considered (John Morse, personal communication, 

10/27/06). Given the recent evidence of the role of P in controlling phytoplankton 

production in the plume and near-plume regions, this process needs further examination. 

 Sulfate reduction is particularly important in affecting the P cycle of coastal 

marine sediments, since it can transform highly adsorptive forms of iron (III) oxides and 

hydroxides into non-sorptive iron (II) sulfides (Krom and Berner, 1980; Caraco et al., 

1989, 1990; Blomqvist et al., 2004). Sulfate reduction may also release P from 

covalently bound minerals as diagenesis proceeds (Sutula et al., 2004). Sulfate reduction 

dominates the metabolism of the sediments to the west of the Mississippi River on the 

Louisiana shelf away from the immediate plume of the river (Rowe et al., 2002), as is 

true for many coastal marine sediments (Howarth, 1984). Sutula et al. (2004) have 

demonstrated that the P content of these sediments is only half that of the riverine 

sediments in the Mississippi from which they are derived due to losses during diagenesis. 

Sulfate reduction and the concomitant changes in sediment iron chemistry may not be the 

only factor involved. Sutula et al. (2004) noted that significant sediment P is lost in the 

immediate plume area of the Mississippi River, a high-energy environment subject to 

physical mixing and sediment reworking, which may make sulfate reduction unlikely [the 

“sub-oxic fluidized bed reactor” processes that Aller (1998) described for other riverine 

plumes]. 

 Studies in the Gulf of Mexico have shown that aerobic respiration in the 

sediments is low during hypoxic events (Rowe et al., 2002). This result suggests that 

anaerobic respiration, the accumulation of reduced compounds, and subsequent oxidation 
48

of these reduced species in the benthic boundary layer (BBL) and sediments may account 

for a large percentage of the oxygen draw down in this area (Morse and Rowe, 1999). 

Other work has found that the balance between the frequency of seabed disturbance, rate 

of geochemical reactions, and reactant concentrations work together to promote efficient 

remineralization through redox cycling in highly mobile muds near large river (McKee et 

al., 2004; Aller et al., 2004; Chen and Gardner, 2004; Chen et al., 2005). This frequent 

cycling of reduced and oxidized compounds is likely to have a profound effect on shortterm oxygen consumption in the BBL, which could influence development of bottom 

hypoxia. 

 Hypoxia and bottom water oxygen deficiency influence not only the habitat of 

living resources but also the biogeochemical processes that control nutrient 

concentrations in the water column. Internal feedbacks on biogeochemical processes 

occur with oxygen depletion. Increased P flux from sediments into overlying waters with 

hypoxia is a classic response in freshwater systems (Mortimer, 1941) and has been welldocumented in coastal marine ecosystems (Nixon et al., 1980; Conley et al., 2002a, 

2002b). However, relatively little work has been done on the Mississippi River shelf on 

estimating the magnitude of enhanced P release with hypoxia and the impact on the 

overall P biogeochemical cycle. Higher P levels do accumulate in the bottom waters of 

the NGOM during hypoxia, but there is no evidence that this mixes into the overlying 

photic zone where it could be available to phytoplankton. This is critical information as 

P can be an important limiting nutrient in the plume (Sylvan et al., 2006). 

 Hypoxia also may influence rates of denitrification. Denitrification is one of the 

major losses of fixed nitrogen in the oceans (Seitzinger and Giblin, 1996), however, its 

measurement is difficult (Groffman et al., 2006). Denitrification is the reductive 

respiration of nitrate or nitrite to N2 or N2O and includes the recently discovered 

anaerobic ammonia oxidation (ANNAMOX) process (Dalsgaard et al., 2003). The rates 

of denitrification are dependent on a variety of factors, but a major control is the 

availability of starting products [e.g. nitrate (Kemp et al., 1990) and carbon (Smith and 

Hollibaugh, 1989; Sloth et al., 1995)]. Note that denitrification is favored by the absence 

of oxygen, but most coastal marine sediments are anoxic below the top few mm. Given 

that large-scale increases in nitrate concentrations and in productivity that have occurred 

on the Mississippi River shelf, it is likely that the rates of denitrification have also 

increased through time. Very few measurements on this important process are available, 

however. 

 An open question is how much hypoxia affects the annual rates of denitrification. 

Few direct measurements of denitrification exist for the Mississippi River shelf, with 

most previous estimates using potential denitrification rates. Lower rates of potential 

denitrification were observed in the Gulf of Mexico zone of hypoxia when low oxygen 

concentrations were encountered (Childs et al., 2002, 2003), although the observed rates 

were at the low end of rates reported for other systems (Herbert, 1999). Denitrification 

can be limited by the availability of nitrate, and hypoxia may reduce the supply rate of 

nitrate by slowing rates of nitrification (the oxidation of ammonium to nitrate); however, 

nitrate concentrations in the hypoxic area were high enough in the Childs et al. (2002) 
49

study not to be limiting. In addition, sulfide, which is commonly found in anoxic 

environments, acts to inhibit nitrification (the oxidation of ammonium to nitrate) (Joye 

and Hollibaugh, 1995), thus reducing the availability of nitrate. In Danish coastal waters, 

rates of denitrification are highest during winter when nitrate concentrations are at their 

annual maximum (Nielsen et al., 1995), and low rates are observed during the summer. 

There are no seasonal measurements of denitrification available for the NGOM to 

estimate the overall effect of hypoxia. In general, the overall rates of denitrification are 

believed to be lower with hypoxia (Sørensen et al., 1987; Graco et al., 2001) and 

eutrophication (Smith and Hollibaugh, 1989), although Vahtera et al. (2007) suggest that 

denitrification has potentially increased with hypoxia. Water column rates of 

denitrification in the oceans are high in mid-water hypoxia areas (Deutsch et al., 2007). 

Further investigations of the effects of hypoxia on the rates of denitrification are sorely 

needed on the Mississippi River shelf, as this is the major pathway of nitrogen loss. 

 Measurement of the fluxes of N and P from sediments provides a direct means to 

assess the role of sediment processes on the relative balance of N and P in the overlying 

water column. There are relatively few NGOM studies where both N and P fluxes from 

sediments have been determined simultaneously. A compilation of these studies shows a 

dissolved inorganic nitrogen/dissolved inorganic phosphorus (DIN:DIP) flux ratio that 

varies from approximately 1:1 to 25:1, with a mean of ~10:1 (Twilley et al., 1999). 

Key Findings and Recommendations 

 The SAB Panel finds that additional information is needed on internal 

biogeochemical processes controlling the availability of nutrients to support primary 

production in the NGOM. The SAB Panel recommends that research be conducted in the 

following areas. 

• The dynamics of sediment/water exchanges of P on the Louisiana shelf and their 

relative role in P cycling. Information on both aerobic and anaerobic processes is 

needed. 

• The effects of hypoxia on the rates of denitrification and on long-term burial and 

regeneration of C, N, and P on the Louisiana shelf. 

• N and P biogeochemical processes in sediments that include analysis of oxygen 

dynamics and the rates of supply of oxygen to the sediment surface. 

2.1.7. Possible Regime Shift in the Gulf of Mexico 

 Hypoxia can act as a positive feedback to enhance the effects of eutrophication 

(Vahtera et al., 2007). It has long been known in lakes (Mortimer, 1941) that the internal 

P loading from sediments during anoxia can sustain eutrophication. In the Baltic Sea, 
50

which is one of the largest coastal areas in the world to suffer from eutrophicationinduced hypoxia, large internal P loading occurs with hypoxia. The amount of DIP 

released from sediments in the Baltic is an order of magnitude larger than external inputs 

from rivers (Conley et al., 2002a). Large sediment-water fluxes of DIP with hypoxia 

must also occur in the Gulf of Mexico, returning DIP to a partially P limited water 

column (Sylvan et al., 2006), stimulating phytoplankton growth and acting as a positive 

feedback to increase hypoxia severity. As discussed earlier (Section 2.1.6), hypoxia has 

the potential to reduce rates of denitrification, which would cause less N to be lost from 

the system, and also act as a positive feedback to increase hypoxia severity. 

 Recent studies in other coastal marine ecosystems, including Chesapeake Bay 

(Hagy et al., 2004) and Danish coastal waters (Conley et al., 2007), suggest that repeated 

hypoxic events can help to sustain hypoxic conditions. Large-scale changes in benthic 

communities occur with hypoxia, reducing the abundance of large, slow growing, deeper 

dwelling animals and facilitating smaller, fast growing species that can colonize surface 

sediments rapidly following hypoxia (Diaz and Rosenberg, 1995). Reductions in the 

abundance and size structure of benthic organisms have been observed in the NGOM 

with hypoxia (Rabalais et al., 2001a). These smaller, surface-dwelling species have less 

capability to irrigate and bring oxygen downward into the sediments, helping to keep the 

sediments anoxic. The loss of benthic communities and the inability of the communities 

to recover with repeated hypoxic events (Karlson et al., 2002) may make ecosystems 

more vulnerable to the development and persistence of hypoxia. In addition, with the 

loss of sediment buffering capacity through the loss of electron acceptors (NO3, O2, Fe2+, 

Mn2+), there is a change in sediment metabolism from aerobic to anaerobic pathways, 

changing the rates and processing of organic matter. 

 Wiseman et al. (1997) showed that the area of hypoxia along the Louisiana-Texas 

shelf was correlated to Mississippi River flow. These relationships were similar to those 

found for Chesapeake Bay (Boicourt, 1992) demonstrating the important role of river 

inputs in providing both freshwater induced stratification and adding nutrients stimulating 

phytoplankton production. However, this apparent relationship has broken down since 

1993 (data provided by DiMarco, personal communication). It appears that the Gulf of 

Mexico hypoxia has worsened following the record breaking 1993 spring floods, e.g., 

smaller river flows now induce a larger response in hypoxia (see Section 2.1.2). The first 

large (>15,000 km2

) hypoxic event occurred after the 1993 flood, with large hypoxic 

areas over 15,000 km2

 observed in most following years. This pattern of a more sensitive 

system is also evident with May-June nitrate loading causing a larger hypoxic area in the 

NGOM than prior to 1993 (data not shown). A similar pattern of an increasingly 

sensitive system following the initial occurrence of hypoxia has been observed in Danish 

coastal waters with worsened hypoxia following the first appearance of large-scale 

hypoxic events (Conley et al., 2007). 

 Changes such as those described above suggest that a regime shift has occurred in 

coastal marine ecosystems that have been affected by large-scale hypoxia (Conley et al., 

2007). Regime shifts are rapid transitions that change the structure and functioning of the 

ecosystem from one state to another as a consequence of a change in an independent 
51

variable. Once a threshold is passed, the ecosystem changes to a new alternative state, 

with changes in biological variables that can propagate through several trophic levels 

(Scheffer et al., 2001; Collie et al., 2004). For example, an increase in certain pelagic 

species (e.g., gelatinous carnivores) can disrupt top-down control of the food web 

structure causing a regime shift to an alternative stable state. The new stable system may 

not respond to changes in nutrient levels, a bottom up control, until nutrient input is 

reduced to a point below which the regime shift occurred. A regime shift due to hypoxia 

implies that, due to hysteresis in the system, nutrients will need to be reduced below the 

level at which the threshold occurred in order to reduce hypoxia. The management 

implications are that nutrients should be reduced as soon as possible before the even 

larger nutrient reductions are required to reduce the area of hypoxia. 

Regime shifts can have large consequences for fisheries (Collie et al., 2004; Oguz 

and Gilbert, 2007). The Gulf of Mexico ecosystem is a tremendously valuable resource 

from economic, ecological and social perspectives. In 2004, the value of commercial fish 

harvest in the Gulf of Mexico was $670 million (NOAA, 2007). The Gulf of Mexico 

shrimp fishery is among the most valuable fisheries in the nation, with a total value in 

2004 of about $370 million, and about $140 million in Louisiana alone. Additionally, an 

estimated 24.6 million recreational fishing days occurred in the Gulf of Mexico in 2004, 

with about 4.8 million of those occurring in Louisiana waters (NOAA, 2007). The Gulf 

of Mexico also serves as habitat for a host of other species, including endangered sea 

turtles and marine mammals. Thus, the Gulf of Mexico is a tremendously valuable 

resource that is potentially being threatened by hypoxia. 

 Earlier studies found it difficult to identify impacts of hypoxia in fisheries 

landings statistics (Diaz and Solow, 1999; Rabalais and Turner, 2001), although there has 

been a shift in relative population abundance from benthic to pelagic species (Chesney 

and Baltz, 2001). A summary of published studies, as well as works in progress, on the 

effects of hypoxia on living resources in the NGOM are mentioned in Appendix A. 

There is strong scientific evidence that ecosystems in the northern Gulf of Mexico are 

stressed by hypoxia (Diaz et al., 2003). Studies have found impacts ranging from the 

molecular/genetic level (Brouwer, 2006; Hendon et al., 2006; Perez et al., 2006; Wells et 

al., 2006), the organismal level (Brouwer, 2006; Zou, 2006) and the ecosystem level 

(Craig et al., 2001; Rabalais and Turner, 2001a; Rabalais, 2006). Potential impacts have 

been identified due to displacement from preferred habitat (Craig and Crowder, 2005; 

Craig et al., 2005; Switzer, et al., 2006). There is also recent evidence that hypoxia has 

affected the valuable brown shrimp fishery (Zimmerman and Nance, 2001). 

 There are some indications that the Gulf of Mexico has undergone a regime 

shift. In the hypoxic/anoxic zone of the Louisiana inner shelf many taxa are lost during 

the peak of hypoxia. Certain typical marine invertebrates are absent from the fauna, for 

example, pericaridean crustaceans, bivalves, gastropods, and ophiuroids (Rabalais and 

Turner, 2001a). As noted above a shift has been observed in the relative abundance of 

fish species. Changes in benthic and fish communities with the change in frequency of 

hypoxia are cause for concern. If actions to control hypoxia are not taken, further 

ecosystem impacts could occur within the NGOM, as has been observed in other 
52

ecosystems. The recovery of hypoxic ecosystems may occur only after long time periods 

(Diaz, 2001) or with further reductions in nutrient inputs. Experience has shown 

recovery to be greatly delayed, taking years to decades for ecosystems to recover after 

nutrient inputs are reduced, and with probably less than complete recovery possible (e.g., 

Diaz, 2001; Diaz et al., 2003; Mee, 2006; Raloff, 2004). Some smaller organisms may 

respondmore rapidly and on annual cycles. For example, in low load years there is less 

hypoxia, lower phytoplankton biomass and presumably less organic deposition and lower 

rates of sediment processes. On the other hand, larger benthic organisms respond more 

slowly, and resident fish and shellfish populations will require more time to return to 

previous conditions. One potential concern with regime shifts is that the condition is not 

always reversible. The system can follow a different path to pre-impact conditions and 

not return to its former state. This is called a hysteresis effect. However, given that the 

Gulf of Mexico is an open shelf system, recovery should be more rapid than in enclosed 

ecosystems. Thus, there are potentially large benefits that justify taking action to control 

hypoxia, and thereby avoiding large-scale changes in the Gulf of Mexico ecosystem. 

Key Findings and Recommendations 

Hypoxia probably increases sediment-water fluxes of P and may reduce the 

potential for denitrification, and change the degradation of organic matter in sediment 

from aerobic to anaerobic metabolism. Biological changes have occurred in the benthic 

communities of the NGOM, and there is evidence that the living resources are impacted 

by hypoxia. The Gulf of Mexico ecosystem appears to have gone through a regime shift 

with hypoxia such that today the system is more sensitive to inputs of nutrients than in 

the past, with nutrient inputs inducing a larger response in hypoxia as shown for other 

coastal marine ecosystems (Chesapeake Bay, Danish coastal waters). The SAB Panel 

therefore provides the following recommendation. 

• Nutrients should be reduced as soon as possible before the system reaches a point 

where even larger reductions are required to reduce the area of hypoxia. 

2.1.8. Single Versus Dual Nutrient Removal Strategies 

 The Action Plan seeks to significantly reduce the size of the Gulf of Mexico 

hypoxic zone by the year 2015, primarily through reductions in nitrogen (N) loadings 

from the MARB to the NGOM. Increases in N loads have clearly been occurring 

throughout the past decades, and there is ample evidence to conclude that N from the 

MARB is a driving force in determining, at least in part, the timing, severity and extent of 

the hypoxic zone. Since the mid-90s, N loadings from the MARB have decreased, 

although they are still much elevated over historic levels. Total phosphorus loadings, 

however, have not changed greatly during this period (Battaglin, 2006; Turner et al., in 

press; Section 2.1.9 of this report). This trend in nutrient loadings has led to reduced 

(albeit still very high by “Redfield” standards) N:P ratios. This evidence suggests that P 
53

is an additional nutrient of concern, in terms of input reductions. As conveyed in 

previous sections of this report, a number of investigators (Sylvan et al., 2006; Dagg et 

al., in press) have concluded that P is limiting primary production during key periods of 

high productivity and in zones of high biomass accumulation in the NGOM adjacent to 

hypoxic waters. Therefore, the role of P in the onset, extent, and duration of the hypoxic 

zone is worthy of additional consideration. 

 Many factors influence the cycling and ultimate fate of both N and P. As both 

play a significant role in driving primary production within the NGOM (and perhaps, in 

conjunction with Si, in the composition of the primary producers and the likely fate of 

produced organic carbon), it is logical to consider the potential for removal of either or 

both elements as a means to reducing hypoxia. The 2001 Action Plan focuses on N 

reductions but does not preclude either P reduction or dual removal strategies. For 

example, the most recent report of the Mississippi River/Gulf of Mexico Watershed 

Nutrient Task Force’s Management Action Review Team (MART, 2006a) concludes that 

most load reduction projects developed under the Clean Water Act Section 319 program 

have targeted both N and P for reduction. Indeed, Howarth et al. (2005) noted that some 

N control practices utilized in the U.S. effectively remove P as well, although the reverse 

is not always the case. However, not all control practices will be effective as a dual 

nutrient removal strategy; see specific discussion on this topic in Section 4.5.10. 

 Restoration plans that focus on N alone may not rapidly improve the situation in 

the MARB where many streams and river segments are degraded by excess P 

concentrations (Action Plan, MR/GMWNTF, 2001). Given recent discoveries 

concerning the importance of P in production of organic carbon within significant 

portions of the NGOM, focusing on N reduction alone may be insufficient to provide the 

desired reduction in the hypoxic zone. However, some plans being undertaken to reduce 

non-point sources of N [forested buffers, 319 programs, and others (see Section 4.4.2, for 

example)] will also lead to P reductions, as well. Reductions in P alone will alleviate 

some of the water quality issues facing freshwater regions of the MARB but are not 

likely, given our current state of understanding, to significantly address the overenrichment of the NGOM. Therefore, greater emphasis on a dual nutrient removal 

strategy is warranted, a conclusion that has been reached in other instances (e.g., National 

Research Council, 2000; Boesch, 2002; Howarth and Marino, 2006). 

 Further work is necessary to examine how effectively current reduction strategies 

target both elements. There may be areas where shifts in removal techniques could 

improve P reduction. In addition, there is still much to be learned about the response of 

autotrophic and microbial communities to shifts in nutrient loading and ratios. A better 

understanding of how these communities have responded to the current loadings and 

predictions of how they will continue to adapt to nutrient reductions will greatly improve 

predictions of the likely response in the extent and duration of hypoxia to nutrient 

reductions in the future. 
54

Key Findings and Recommendations

 Recent information clearly indicates that P controls productivity in some portions 

of the NGOM. The SAB Panel finds that restoration plans focusing on N alone may not 

rapidly improve the situation in the MARB and may be insufficient to provide the desired 

reduction in the hypoxic zone. Reductions in P alone will alleviate some of the water 

quality issues facing freshwater regions of the basin but are not likely to significantly 

address the over-enrichment of the NGOM. Therefore the SAB Panel offers this 

recommendation. 

• In addition to the N reduction strategy currently in place, reduction strategies for 

P should be implemented. Section 4.2 provides greater detail on the SAB Panel’s 

recommended targets for reducing both N and P. 

2.1.9. Current State of Forecasting 

 There are several types of modeling efforts working toward a better understanding 

of factors influencing the extent and duration of the Gulf of Mexico hypoxic zone. These 

vary from the simple to the complex and are based on empirically observed relationships, 

on mechanistic understanding, or some combination of both. 

 Empirical models are widely used in the aquatic sciences to establish relationships 

between variables, with the most well known being the correlation between spring P 

loading in lakes and summer chlorophyll concentrations (Vollenweider, 1976). This 

work has been widely used in a management context to justify reductions in 

anthropogenic phosphorus loading to lakes and to set goals for reductions for particular 

lakes. Nixon et al. (1996) developed a similar correlation between annual loading of DIN 

and rates of primary productivity for marine ecosystems. While establishment of 

empirical models has greatly enhanced understanding of the structure and functioning of 

aquatic ecosystems (Peters, 1986), the standard criticism of this approach is that 

correlation does not imply causation. Although correlations between variables exist, they 

do not explain why variables are correlated or the mechanisms of the relationship. They 

do, however, provide some very useful predictive capability. In addition, when 

ecosystem production is greatly different from that predicted, controls on productivity 

other than nutrients may be dominating, such as light limitation or limitation from rapid 

flushing (Howarth et al., 2006a). 

 Some new forecast modeling work has been completed since the Integrated 

Assessment. Turner et al. (2006) developed simple linear and multiple regression models 

to examine hypoxia in the NGOM. Empirical models require important decisions 

regarding the choice of variables and of the time scales of model operation. Turner et al. 

(2006) tested many different nutrient loading lag times and concluded that the best 
55

relationship was obtained two months (May) prior to the maximum observed extent of 

hypoxia (July), with significant correlations for nitrate+nitrite, total nitrogen (TN), orthoP and total phosphorus (TP) (r2

 values of 0.50, 0.27, 0.54, and 0.60, respectively). A 

multiple regression analysis was also developed incorporating nutrient load and a new 

variable “Year” to account for the increase in carbon in surface sediments after the 1970s 

causing significantly more sediment oxygen demand. A lag of two months of nutrient 

loading was, again, the most significant variable to describe hypoxic area with r2

 values 

of 0.82, 0.80, 0.69, and 0.64 obtained with nitrate+nitrite, TN, ortho-P and TP, 

respectively. Turner et al. (2006) then used the nitrate+nitrite model to extrapolate 

beyond the data range used to construct their models to predict hypoxic area prior to 

available measurements. When the hindcasted values became negative, they were plotted 

as zero values. In general, it is considered incorrect to extrapolate model results in this 

manner beyond the range of the data supporting the model, as other mechanisms and 

relationships may exist that may not be included in the regression analysis. Further, the 

SAB Panel believes that the addition of the variable “Year” in the multiple regression 

analysis is inappropriate as the addition of one more year will cause prediction of a 

positive increase in hypoxia with time. 

 Among models that address Gulf of Mexico hypoxia and include some 

consideration of processes and mechanisms, that of Scavia et al. (2003) is one of the 

simplest. Their model uses a relationship between the nitrogen loading from the MARB 

and the decay of oxygen “downstream” (i.e., in the NGOM - within the plume and the 

nearshore reaches to the west of the Mississippi and Atchafalaya River outflows). When 

used in a forecast mode, this model is able to only explain approximately 45-55% of the 

variability in hypoxic length and area. This model explicitly addressed uncertainty in 

prediction. The SAB Panel found this approach to be very useful. Recently, in 

combination with a watershed model, the model of Scavia et al. (2003) has been used to 

address how climatic variability and change may affect Gulf hypoxia (Donner and 

Scavia, 2007). A similar model has also been applied very successfully to understand 

hypoxia and anoxia in Chesapeake Bay (Scavia et al., 2006). The Scavia et al. (2003) 

model focused on N loading and did not consider P. Consideration of P would seem to 

be a timely addition to the model, and a manuscript including P recently was accepted for 

publication by Scavia and Donnelly (Scavia and Donnelly, in press). This model 

approach, and the modeling efforts of Bierman and colleagues and Justic and colleagues 

(see below) all provide reasonably consistent guidance and suggest similar levels of N 

reduction that might be required to reduce the extent of the hypoxic zone. 

 Other process-based models are more complex and attempt to model both 

physical and biological controls occurring in the hypoxic region. Examples include those 

of Bierman et al. (1994), Justic et al. (1996, 2002), and Green et al. (2006b). The 

Bierman et al. (1994) model is the most complex of these approaches and simulates the 

steady-state summertime conditions for the hypoxic area using three-dimensional 

modeling of the physics as well as interactions between food web processes, nutrients, 

and oxygen. The model of Justic et al. (1996, 2002) simulates oxygen dynamics at one 

location within the hypoxic zone using a simple model that has two vertical layers and 

meteorological conditions and nitrogen loads as drivers. The Green et al. (2006b) surface 
56

mixed layer model is based on food web dynamics and relatively simple two-dimensional 

physics (no vertical dimensionality) of the Mississippi River plume. This model predicts, 

among other things, the relationship between carbon sources and bottom-water oxygen 

depletion; the model does not include changes to either N or P inputs or dynamics. None 

of these more complex models explicitly presented analysis of uncertainty or sensitivity 

analysis of potential biasing terms. As with the Scavia et al. (2003) model, Bierman et al. 

(1994) and Justic et al. (1996, 2002) do not consider P loads or dynamics. 

 It should be pointed out that complex water quality models that could be very 

useful in the NGOM have been developed and used in other environmentally stressed 

regions like the Chesapeake Bay system (Cerco and Cole, 1993), Long Island Sound (St. 

John et al., 2007), the New York/New Jersey Harbor/New York Bight complex 

(Landeck-Miller and St. John, 2006), and the Massachusetts/Cape Cod Bays system 

(Besitkepe et al. 2003). These models include a coupling to three-dimensional and timedependent hydrodynamics, a water column eutrophication submodel and a sediment 

diagenesis/nutrient flux submodel. The water column eutrophication submodel includes 

state-variables for three functional phytoplankton groups; dissolved inorganic nutrients 

(ammonium, nitrate+nitrite, ortho-phosphate, and silica);, and labile and refractory forms 

of dissolved and particulate organic nitrogen and phosphorus, biogenic silica; labile and 

refractory forms of particulate and dissolved organic carbon; and dissolved oxygen. The 

sediment nutrient flux submodel includes state-variables for labile, refractory, and inert 

organic carbon, nitrogen, and phosphorus, as well as biogenic silica. Inorganic 

substances tracked include ammonium, nitrate+nitrite, ortho-phosphate, silica, sulfide, 

and methane. Processes tracked in the sediment flux model include: organic matter 

deposition; sediment diagenesis; burial; the flux of inorganic nutrients between the water 

column and the sediment bed; and the generation of sediment oxygen demand (SOD). 

There is an inherent trade off between model simplicity (where many potentially 

important factors are not considered) and complexity (where many coefficients and a 

great amount of data are required). More complex models may have value to help devise 

effective management strategies, especially if N reductions alone will not be sufficient to 

control hypoxia and if the more complex models can reasonably capture the importance 

of P. However, with complexity comes greater numbers of estimated parameters and the 

uncertainty associated with them. Hence this type of model may not improve forecasting 

capabilities dramatically. The development of more complex models is likely to prove 

extremely valuable for understanding the physical factors controlling water and carbon 

(C) transport, the dynamics of nutrient interactions with primary producers, and the 

recycling and loss of C and nutrients from the system. There is also great value in 

refining and further developing simple models, which may, in the end, prove most 

valuable for making management decisions. Scavia et al. (2004) explicitly compared the 

models of Scavia et al. (2003), Biermann et al. (1994), and Justic et al. (1996, 2002) for 

use in managing Gulf of Mexico hypoxia and showed that all three models gave broadly 

consistent guidance. 

 The physics of the NGOM region is complex, and there is clear value in 

developing more complex models of physical processes for this region. Improved three-
57

dimensional models with finer grid structure than present models would have many uses. 

These uses include assisting the interpretation of monitoring data and serving as 

platforms upon which improved models of biogeochemistry and ecological response 

could be built. However, the level of complexity in the biogeochemistry and ecology 

need not match the complexity of the physical models (Hetland and DiMarco, 2007). 

Complex physical models could be very valuable in constructing simple box massbalance accounting models for C, N, P, Si, and O, for example. The importance of 

developing such budget-based models is discussed further below. 

 In addition to statistical and simulation models, another modeling format that 

should be considered involves construction and evaluation of material budgets or mass 

balance models. These are basically quantitative input-output budgets with additional 

complexity added by consideration of internal processes of production, recycling and 

loss. These relatively simple budgets provide a quantitative mass balance framework to 

test the understanding of how the systems work. These budgets should be developed on a 

seasonal basis (e.g., summer hypoxic season) and evaluated for distinctive areas (e.g., 

Mississippi River Plume). These budgets are largely based on empirical observations and 

are not simulated through time, although data used in a budget analysis are needed in 

simulation models for both calibration and verification. As an example, an oxygen 

budget (Equation 1) would involve DO inputs/outputs from air-sea diffusion, horizontal 

advective/dispersive transport, and vertical transport between euphotic and subpycnocline zones. In addition, DO is added through daytime photosynthesis and lost 

through water column and sediment respiration. Evaluation of these pathways indicates 

especially important processes, and imbalances in the budget point to areas where 

understanding or measurements are inadequate. We suggest that conceptual mass 

balance models also be used to provide a checklist of needed measurements for future 

NGOM hypoxia research/monitoring.

 Other general points regarding modeling efforts are summarized in Section 3.4 of 

this report. An important conclusion for both models of the response of the NGOM to 

nutrient inputs and watershed models generating estimates of nutrient loads is that a 

diverse ensemble of models is needed, including both relatively simple and more 

complex ones. No one best approach to modeling can be identified, and management of 

Gulf hypoxia is best served by having multiple models with multiple outputs. The SAB 

Panel suggests that modeling efforts, ranging from the simple to complex, be conducted 

in parallel wherein there is the opportunity for cross-testing of results among model 

formats. When predictions tend to agree, managers can have more confidence in 

deciding upon courses of action. When models do not agree, dissecting the reasons for 

divergence can lead to better understanding and, ultimately, better management. 
58

Key Findings and Recommendations

 Since the Integrated Assessment, a number of modeling approaches have been 

employed to characterize the onset, volume, extent, and duration of the hypoxic zone. 

Models have been able to explain approximately 45-55% of the variability in hypoxic 

length and area. However, the SAB Panel finds that model development, calibration, and 

verification are hampered by the relative paucity of data on the duration and extent of 

hypoxia and on rates of important biogeochemical and physical processes that regulate 

hypoxia. In addition, the SAB Panel finds that a diverse ensemble of models is needed, 

including both relatively simple and more complex ones. No one best approach to 

modeling can be identified, and management of Gulf hypoxia is best served by having 

multiple models with multiple outputs. The SAB Panel provides the following 

recommendations to advance the science for characterizing the onset, volume, extent, and 

duration of the hypoxic zone. 

• To the extent reasonable, future models (particularly more complex models that 

rely on accurate representation of ecological and biogeochemical processes) of 

hypoxia in the Gulf should consider nitrogen, phosphorus, and their interactions. 

However, this is a significant challenge since these interactions are so poorly 

studied in the NGOM at present. 

• The development of more comprehensive monitoring should be coordinated with 

model development. For example, the more complex physical models of the 

NGOM should be used to aid in interpretation of monitoring data on extent and 

duration of hypoxia. These models can also feed into both simple and complex 

biogeochemical and ecological models. 

• Because there is great value in developing simple mass balance models in the 

NGOM for organic C, dissolved oxygen, and nutrients, mass balance models 

should be used to provide a checklist of needed measurements for future NGOM 

hypoxia research/monitoring. 

• Gulf hypoxia models should be designed so that they can be compatible with 

watershed models. That is, there must be compatibility in 1) the time step 

between a Gulf hypoxia model and a watershed model, and 2) the form of key 

variables that serve as outputs from a watershed model and inputs for a Gulf 

hypoxia model (e.g., a watershed model that predicts total nitrogen is not 

compatible with a Gulf hypoxia model that requires specific forms of nitrogen). 
59

3. Nutrient Fate, Transport, and Sources 

 The SAB Panel was asked to review the available literature and information, 

especially that developed since 2000, that would allow them to assess any changes and 

improvements in the understanding of nutrient sources and flux estimates within the 

Mississippi and Atchafalaya River basins (MARB) (see Figure 2) and the current ability 

to use watershed models to route and predict nutrient delivery to the Gulf of Mexico. 

The following sections discuss the current levels of understanding and provide brief 

summaries of the SAB Panel’s key findings and recommendations. 

3.1. Temporal Characteristics of Streamflow and Nutrient Flux 

 The research needs identified in the Integrated Assessment to understand and 

document the temporal characteristics of MARB riverine nutrient loads included 1) 

studies on small watersheds to better document nutrient export on the short time scales 

needed; 2) detailed information on tile drainage intensity; 3) increased monitoring of 

stream sites; and 4) measurements of point source discharges rather than estimates from 

permits. Only a limited number of these needs have been met. 

 However, more recent estimates of agricultural drainage appear to be more 

representative than those used in the original assessment, and new procedures for load 

calculations have resulted in changes in estimates of nutrient fluxes. A brief discussion 

of each of the improvements follows. 

Current Extent and Patterns of Agricultural Drainage 

 The Integrated Assessment relied largely on the 1987 USDA-ERS report (Pavelis, 

1987), which based estimates of agricultural drainage on land capability class and crop 

information from the 1982 Natural Resources Inventory (NRI). NRI estimates were 

dropped after 1992, and NRI is statistically valid only at a watershed or county level. 

Based on the USDA surveys, some degree of subsurface drainage is present on 13 million 

hectares (over 32 million acres) in the Midwest states. However, there is considerable 

uncertainty with respect to the actual extent and distribution of drainage of cultivated 

cropland. In the absence of additional survey data, more recent estimates of the extent of 

drained agricultural land have been developed based on land use and soil 

class/characteristics (Jaynes and James, 2007; Sugg, 2007). This general approach needs 

further development and validation but seems to provide the best current estimate of the 

extent of agricultural drainage. The approach takes advantage of the now extensive and 

detailed GIS coverages and provides a considerably finer level of spatial resolution than 

previously available. 

 In the following example, USDA STATSGO soil data were used to estimate the 

extent of agricultural drainage based on the distribution of row crops (primarily corn and 

soybean) on soils with a drainage class of poorly drained soils and slopes 2% or less 
60

(Figure 10, per D. Jaynes, National Soil Tilth Lab, Ames, IA). The patterns of 

agricultural drainage predicted using this approach are generally similar to patterns in 

land use (Figure 11) and in-stream nitrate concentration estimated from STORET data 

selected to exclude point source influences (Figure 12). Drainage estimates could be 

further refined by using improved land use data and by using SSURGO rather than 

STATSGO data. 

Figure 10: Estimated extent of agricultural drainage based on the distribution of row crops, largely corn 

and soybean, and poorly drained soils (per D. Jaynes, National Soil Tilth Lab, Ames, IA). 
61

Figure 11: Land cover based on Landsat data (adapted from Crumpton et al., 2006). 

Figure 12: Flow weighted average nitrate concentrations estimated from STORET data selected to exclude 

point source influences (adapted from Crumpton et al., 2006). 

 The relationship between nitrate concentration and land use is further illustrated 

in Figure 13 for 52 NASQAN stations (Alexander et al., 1998) in the upper Mississippi 

and Ohio River basins selected to exclude sites with large upstream reservoirs or 
62

extensive upstream urban areas (Crumpton et al., 2006). See Section 4.5.7 for further 

discussion on urban non-point sources. Percent cropland (corn or soybean) accounts for 

90% of the observed variation in the average of 1980 to 1993 annual flow-weighted 

average nitrate concentrations for the 52 stations examined. Reduced nitrogen 

(calculated as total nitrogen minus nitrate) shows a slight, but statistically significant, 

increase with percent crop land. 

0 10 20 30 40 50 60 70 80 90 100

Percent crop land

0

2

4

6

8

10

12

14

Flow weighted concentration

(mg N L-1)

Nitrate

Reduced N (TN-nitrate)

Model nitrate concentration (R2 = 0.90)

Linear fit to reduced N data

n = 52

Figure 13: Flow-weighted average nitrate and reduced N versus percent cropland (adapted from Crumpton 

et al., 2006).

 Flow-weighted average nitrate concentrations estimated by applying the 

regression for NASQAN sites to 1992 Landsat land cover data for UMR and Ohio River 

basins are similar to those estimated from STORET data. The relationship between 

nitrate concentration and the estimated extent of agricultural drainage was also examined, 

and for these 52 stations, nitrate concentrations were more closely related to land use than 

to STATSGO derived estimates of drainage. There is certainly more error in estimates of 

drainage than in estimates of cropland distribution, and this error could degrade the fit of 

nitrate concentration with drainage. However, much of the cropland not directly drained 

by field tile still contributes to nitrate discharged through drainage networks, and at some 

spatial scale, nitrate concentrations might depend more on cropland distribution than on 

artificial drainage (i.e., if the land is successfully cropped, then some combination of 

natural and artificial drainage can be implied). 

 It is clear that agricultural drainage in the Corn Belt is extensive, the general 

distributions of drainage and cropland are correlated, and nitrate concentrations are 

correlated with patterns of cropland and drainage. Additional research is needed to better 

define the extent, pattern, and intensity of agricultural drainage, including cropland 

drained by field tile as well as cropland not directly drained by field tile but contributing 

to drainage networks. 
63

Change in the Flux Estimation Method 

 Riverine loads can be calculated with many different methods; the method chosen 

is dependent on sampling frequency as well as river size, which determines how quickly 

the concentration changes. A comparison of the estimates of annual N flux for the 

combined Mississippi and Atchafalaya Rivers using five different methods is shown in 

Figure 14. Goolsby et al. (1999) presented nitrate-N loads to the Gulf for 1955 through 

1996. For the years prior to 1968, loads were calculated from either daily samples 

composited at 10- to 30-day intervals for analysis. For the period 1968-1996, they used a 

multiple regression approach to calculate daily concentrations based on about 10 to 15 

samples per year (or less) and daily flow [shown as Goolsby et al. (1999)]. Goolsby et al. 

(1999) calibrated one model (using a minimum variance unbiased estimator, MVUE) for 

1968-1975, and one model for 1976-1997. This type of regression equation provides a 

good measure of the overall flux of a nutrient for the entire period of fitting but is less 

accurate for a given year. Since the Integrated Assessment, USGS has modified load 

estimation procedures to reduce the bias in the regression models. These modified 

procedures are all based on the rating curve method but differ in the form of the equation 

and/or calibration periods. In July 2002, USGS posted load estimates for the entire 

period of record using ESTIMATOR (Cohn et al., 1992; Gilroy et al., 1990), a 

regression-model method using the same MVUE technique used by Goolsby et al. (1999) 

with a 10-year moving window calibration period, and provided updated annual estimates 

through June 2002, followed by annual updates through June 2005 (shown as LOADEST 

10 yr). In this case the MVUE procedure used was equivalent to the adjusted maximum 

likelihood estimate (AMLE, discussed below) used in later estimates because there were 

no censored nitrate values in the calibration datasets. In 2006, the USGS posted new 

estimates for the entire period of record using Load Estimator (LOADEST) (Runkel and 

others, 2004) with the AMLE procedure and a 5-year moving window (shown as 

LOADEST 5 yr). In addition to a shorter calibration period, the AMLE procedure 

modifies the rating curve equation in an attempt to correct for transformation bias. 

However, the AMLE procedure can still suffer from serial correlation in the residuals; so 

when sufficient data are available, the USGS applies a period-weighted interpolation to 

correct the AMLE estimate for the serial structure in the residuals (Aulenbach and 

Hooper 2006). Results from this composite method for the mainstem Mississippi and 

Atchafalaya Rivers are nearly the same as just using a period-weighted (or linear 

interpolation) approach for nitrate-N (shown as Composite). This suggests that the 

regression model in the composite method adds little when at least 10 samples are 

available for a given year, as well as demonstrating that concentrations of nitrate-N 

change slowly in these large rivers. (For additional information on methods used to 

estimate nutrient fluxes see: http://toxics.usgs.gov/pubs/of-2007-1080/methods.html.) 
64

1950 1960 1970 1980 1990 2000 2010

Riverine Nitrate-N Flux 

(million tons N yr-1) 0.0

0.5

1.0

1.5

2.0

Composite sampling prior to 1968

LOADEST (5 yr, 1979 on)

Goolsby et al.

LOADEST 10 yr

Composite

Figure 14: MARB nitrate-N fluxes for 1955 through 2005 water years comparing estimates from various 

methods for 1979 to 2005. Based on USGS data from Battaglin (2006) and Aulenbach et al. (2007). 

 Although the overall year-to-year pattern of N flux is consistent across the various 

methods, there is considerable variability amongst the estimates of each annual N flux. 

Figure 15 shows the percent difference between three of the methods and the current 

LOADEST 5 yr method in both percent and metric tons for the entire period of record. 

The LOADEST 10 yr method estimated N fluxes that ranged from as much as about 18% 

less (1990) to 28% more (1994) than the N fluxes estimated by the LOADEST 5 yr 

method. That translates into an underestimate of about 180,000 metric tonne or 198,000 

ton of N that was delivered to the Gulf in 1990 and an overestimate of about 260,000 

metric tonne of N (287,000 ton of N) in 1994. Research published since 2003 would 

have used the LOADEST 10 yr fluxes in models predicting the Gulf hypoxic zone in 

which case they likely used the more recent estimates (2003 and 2004 in Figure 14), 

which ranged from only 3-10% or 25-50,000 metric tonne of N (28-55,000 ton of N) 

more than the estimated flux using the current LOADEST 5 yr method. The flux 

estimates presented in the following sections of this report are based on the new 

LOADEST 5 yr method. 
65

Riverine Nitrate-N Flux Difference

 from LOADEST 5 yr (%)

-30

-20

-10

0

10

20

30

Goolsby et al. 

LOADEST (10 year) 

Composite

1975 1980 1985 1990 1995 2000 2005

Riverine Nitrate-N Flux Difference

 from LOADEST 5 yr 

(million metric tons N)

-0.2

-0.1

0.0

0.1

0.2

0.3

Figure 15: Comparison (percent and absolute basis) of MARB nitrate-N fluxes to LOADEST 5 yr method 

for 1979 through 2005 water years. Based on USGS data from Battaglin (2006) and Aulenbach et al. 

(2007). 

3.1.1. MARB Annual and Seasonal Fluxes 

 The following analysis is based on U.S. Geological Survey streamflow and waterquality monitoring data described in Aulenbach et al. (2007) and available on the internet 

at: http://toxics.usgs.gov/pubs/of-2007-1080/. The nutrient flux estimates were 

calculated as the combined fluxes at the Mississippi River near St. Francisville, LA and 

the Atchafalaya River at Melville, LA (Figure 16) using the LOADEST 5 yr method 

discussed in the previous section. 
66

Figure 16: Schematic showing locations of MARB monitoring sites (Aulenbach et al., 2007).

Annual Patterns 

 Nitrogen -- During the last five years (2001 to 2005 water years), an average of 

813,000 metric tonne (896,000 ton) of nitrate-N and 429,000 metric tonne (473,000 ton) 

of total Kjeldahl N (TKN) were transported annually to the Gulf. There is considerable 

inter-annual variability in these flux values, driven primarily by precipitation patterns and 

resulting streamflow (Figure 17), which appears to have increased slightly since the 

1950s. Since the mid-1990s, annual nitrate-N flux has steadily decreased, which is more 

clearly shown by the 5-year running average. In addition, TKN has also shown a steady 

decline since the mid 1980s, so the total N flux, although highly variable from year to 

year, shows a very striking decline. The annual NH4-N flux also decreased during the 

monitoring period (from 77,000 metric tons N/yr [85,000 tons N/yr] in 1980 to 1984 to 

12,000 metric tons N/yr [13,000 tons N/yr] for 2001 to 2005) but was not the primary 

reason for the decline in TKN, as particulate and organic N declined. The decline in 

NH4-N is likely due to improvements in sewage treatment as is at least part of the decline 

in particulate and organic N (Larson, 2001; Metropolitan Council, 2004). In addition, 

reduced sediment loads, because of a reduction in soil erosion, may also be a driving 

factor in reducing particulate N losses (Richards and Baker, 2002). 
67

Water Flux (million m3

)

0

200000

400000

600000

800000

1000000

1200000

Riverine N Flux

(million metric tons N yr-1) 0.0

0.5

1.0

1.5

2.0

2.5

Annual nitrate

Nitrate five year running average

1955 1960 1965 1970 1975 1980 1985 1990 1995 2000 2005

0.0

0.5

1.0

1.5

2.0

2.5

Ammonium-N

Particulate/Organic N

Total N 

Figure 17: Flow and available nitrogen monitoring data for the MARB for 1955 through 2005 water years. 

(LOWESS, Locally Weighted Scatterplot Smooth, curves shown in red). LOWESS describes the 

relationship between Y and X without assuming linearity or normality of residuals, and is a robust 

description of the data pattern (Helsel and Hirsch, 2002).

Phosphorus and Silicate -- Temporal trends in total P, soluble reactive P (SRP), 

and dissolved silicate fluxes for the combined rivers are less striking than the trends in N 

flux. The average annual total P flux (Figure 18) was 154,000 metric tons P/yr (170,000 

tons P/yr) for the water years 2001 - 2005, with SRP flux 24% of total P flux. Battaglin 

(2006) reported that total P flux increased during that period, but this was in comparison 

to the average flux during the period 1980 - 1996. When total P flux is viewed during the 
68

entire period of 1980 – 2005 and a LOWESSS curve fit to the dataset, there appears to be 

a slight increasing trend since the mid 1990s. The annual flux of dissolved silicate 

appears to have declined slightly since the early 1990s. Water Flux (million m3

)

0

200000

400000

600000

800000

1000000

1200000

Riverine P Flux 

(million metric tons yr-1) 0.00

0.05

0.10

0.15

0.20

Total P

Soluble reactive P

1955 1960 1965 1970 1975 1980 1985 1990 1995 2000 2005

Riverine Silicate Flux 

(million metric tons Si yr-1) 0

1

2

3

4

Figure 18: Flow, available phosphorus, and available silicate monitoring data for the MARB for 1955 

through 2005 water years. (LOWESS curves shown in red). Based on USGS data from Battaglin (2006) 

and Aulenbach et al. (2007). 

 Nutrient Ratios – Ratios of N to P and Si to N can be important in determining the 

growth of various phytoplankton species in the Gulf. The Si:DIN (dissolved inorganic N) 

ratio ranged from about 2 to 4.5 during the 1950s and 1960s but then greatly decreased as 

silicate concentrations declined by about 50% between the 1950s and 1980s (Turner and 

Rabalais, 1991; Rabalais et al., 1999). Ratios since 1980 of Si:DIN have been just above 

1 annually (Figure 19), averaging 1.08 for 2001 to 2005 water years. Nitrogen to P ratios 

averaged 18 for 2001 to 2005 have shown little variability since the early 1990s, with 
69

perhaps a declining trend. These ratios are useful to compare to the Redfield ratio (Si:N:P 

= 16:16:1) and suggest, as Rabalais et al.(1999) concluded, that annual nutrient fluxes to 

the Gulf are quite close to this ratio. However, spring ratios, discussed later, are 

somewhat different and may have a more important effect on Gulf phytoplankton growth. Total N:Total P (mol mol-1) 10

20

30

40

50

60

1980 1985 1990 1995 2000 2005 2010

Si:DIN (mol mol-1) 0.50

0.75

1.00

1.25

1.50

Figure 19: Ratio of total N to total P and dissolved silicate to dissolved inorganic N for MARB for the 

1980 through 2005 water years. Based on USGS data from Battaglin (2006) and Aulenbach et al. (2007). 

Seasonal Patterns 

 Nitrogen -- Since the Integrated Assessment, greater emphasis has been placed on 

the spring flux of nutrients (sum of April, May, and June fluxes) as a possible important 

regulator of hypoxia, and, therefore, fluxes for this period were examined using the 

available data for the period 1979 - 2006. Whereas the annual water flux showed a 

slightly increasing trend since 1990 (Figure 17), the spring water flux, although highly 

variable, appears to show a decreasing trend (Figure 20). Spring nitrate-N flux also has 

declined, with even larger decreases in TKN flux and, therefore, total N flux. 
70

Water Flux (million m3

)

0

100000

200000

300000

400000 Riverine N Flux (million metric tons N)

0.0

0.2

0.4

0.6

0.8

1.0

1975 1980 1985 1990 1995 2000 2005 2010

0.0

0.2

0.4

0.6

0.8

1.0

Total Kjeldahl N

Total N

Nitrate-N

Figure 20: Flow and nitrogen flux for the MARB during spring (April, May, and June) for the period 1979-

2005. (LOWESS curve shown in red). Based on USGS data from Battaglin (2006) and Aulenbach et al. 

(2007). 

 Phosphorus and Silicate -- Spring P flux (both total and SRP) has changed 

relatively little, with perhaps a small decrease in total P flux (Figure 21). The spring 

dissolved silicate flux has shown a pronounced decline since 1990s, greater than the 

decline in water flux. The reason for this decline is not known. 
71

Water Flux (million m3

)

0

100000

200000

300000

400000

Riverine P Flux 

(million metric tons P)

0.00

0.02

0.04

0.06

0.08

Soluble reactive P

Total P

1975 1980 1985 1990 1995 2000 2005 2010

Riverine Silicate Flux

(million metric tons Si)

0.2

0.4

0.6

0.8

1.0

1.2

Figure 21: Flow, phosphorus, and silicate flux for the MARB during spring (April, May, and June) for the 

period 1979-2006. (LOWESS curve shown in red). Based on USGS data from Battaglin (2006) and 

Aulenbach et al. (2007). 

 Figure 22 shows the spring fluxes (sums of April, May, and June fluxes) as a 

percentage of the annual fluxes. There is considerable inter-annual variability in the 

annual fluxes that occurs during spring, as indicated by the whiskers on the box plots. 

Spring water flux was, on average, 30% of annual flux, whereas nitrate-N was 40%, TKN 

34%, and total P 34% of their annual fluxes. Therefore, the river is disproportionately 

enriched with all nutrients during the spring but particularly with nitrate. This result 
72

further substantiates the conclusion drawn earlier that tile-drained fields are a primary 

source of N, which is released beginning in winter (Ohio into central Illinois) to spring 

(northern Illinois, Iowa and Minnesota). This influence was very evident in 2002, when 

50% of the nitrate-N flux occurred during the 3 spring months. Royer et al. (2006) 

pointed out how most of the N and P flux from tile drained watersheds occurred during a 

few months during winter and spring each year, further supporting the trends at this 

larger scale. 

Water Nitrate-N TKN Total P Silicate

Sum of April, May, June Flux

(% of Annual)

0

10

20

30

40

50

60

Figure 22: Sum of April, May and June fluxes as a percent of annual (water year basis) for combined 

Mississippi mainstem and Atchafalaya River. Box plots show median (line in center of box), 25th and 75th 

percentiles (bottom and top of box, respectively), 10th and 90th percentiles (bottom and top error bars, 

respectively) and values < 10th percentile and > 90th percentile (solid circles below and above error bars, 

respectively). Based on USGS data from Battaglin (2006) and Aulenbach et al. (2007). 

 Nutrient Ratios – N to P ratios during spring flow to the Gulf averaged 22 for 

2001 to 2005 (Figure 23), greater than the annual value of 18 for the same time period. 

As discussed previously, nitrate transport is greater during this period than is P transport. 

The Si:DIN ratio was also lower during the spring compared to the annual mean for 2001 

to 2005 (spring ratio 0.84, annual ratio 1.08), reflecting greater transport of nitrate 

compared to silicate. Turner et al. (1999) concluded that decreasing Si:DIN ratios to less 

than 1.1 could greatly alter Gulf food web dynamics because the proportion of diatoms in 

the phytoplankton community would be reduced, which would impact zooplankton and 

higher trophic levels. 
73

Total N:Total P 

(mol mol-1) 10

20

30

40

50

60

1980 1985 1990 1995 2000 2005 2010

Si:DIN (mol mol-1) 0.50

0.75

1.00

1.25

1.50

Figure 23: Ratio of total N to total P and silicate to dissolved inorganic N for the MARB during spring 

(April, May, and June) for the period 1980-2006. Based on USGS data from Battaglin (2006) and 

Aulenbach et al. (2007). 

3.1.2. Subbasin Annual and Seasonal Flux 

Annual Patterns 

 USGS estimates (Aulenbach et al, 2007) were used to examine nutrient fluxes 

within subbasins of the MARB. Annual nutrient fluxes were calculated with an adjusted 

maximum likelihood estimate (AMLE), a type of regression-model method, with a 5-year 

moving average calibration period (composite method estimates were not made for 

subbasin data). Figure 24 shows the location of nine subbasins comprising the MARB 

and Table 2 lists site name and map number for the associated monitoring sites. Figure 

16 shows a schematic of the MARB sampling stations to assist with the following 

analyses. The initial analysis discusses the cumulative fluxes of five major subbasins: 1) 

Upper Mississippi (upstream of Thebes, IL minus the inflow from the Missouri River), 2) 

Ohio-Tennessee (upstream of Grand Chain, IL), 3) Missouri (upstream of Hermann, 
74

MO), 4) Arkansas-Red (combined flux from the Arkansas and Red Rivers), and 5) Lower 

Mississippi. 

Figure 24: Location of nine large subbasins comprising the MARB that are used for estimating nutrient 

fluxes (from Aulenbach et al., 2007). 
75

Table 2: Site name and corresponding map number for sites discussed in the following section.

Site name Map Number 

Mississippi River at Clinton, Iowa 1 

Mississippi River below Grafton, Illinois 2 

Missouri River at Omaha, Nebraska 3 

Missouri River at Hermann, Missouri 4 

Mississippi River at Thebes, Illinois 5 

Ohio River at Cannelton Dan at Cannelton, Indiana 6 

Ohio River at Dam 53 near Grand Chain, Illinois 7 

Arkansas River below Little Rock, Arkansas 8 

Red River at Alexandria, Louisiana 9 

 Annual Flux Estimates: The flux estimates from the five subbasins are listed in 

Table 3. During the last five water years, most of the nitrate-N flux (84%) and TKN flux 

(73%) was from the Upper Mississippi and Ohio-Tennessee subbasins. The Missouri 

subbasin contributed 9.8% of the nitrate-N flux to the Gulf, with much smaller fluxes 

coming from the Arkansas-Red and lower Mississippi River subbasins. These data 

clearly illustrate that the source of both nitrate-N and TKN is from the upper Mississippi 

River basin before the Missouri River enters. For total P flux, the Missouri subbasin was 

more important and contributed 20% of the flux, compared to 26% and 38% for the upper 

Mississippi and the Ohio-Tennessee subbasins, respectively. 

Table 3: Average annual nutrient fluxes for the five large subbasins in the MARB for the 2001-2005 water 

years. (Percent of total basin flux shown in parentheses.) 

Subbasin Area Nitrate-N TKN Total P 

(km2

) 

Flow 

(M m3/yr) (in 1,000 metric tons) 

Upper Mississippi1 493,900 116,200 349 (43%) 136 (32%) 40.4 (26%)

Ohio-Tennessee 525,800 279,800 335 (41%) 175 (41%) 58.7 (38%)

Missouri 1,353,300 60,080 78.6 (9.8%) 83.8 (20%) 30.4 (20%)

Arkansas-Red 584,100 67,200 28.7 (3.5%) 43.9 (10%) 8.7 (6%)

Lower Mississippi1 183,200 129,550 22.1 (2.7%) -8.4 (-2%) 16.1 (10%) 1

Nutrient fluxes calculated by difference. Negative values occur where downstream site had a lower flux 

than upstream site, the result of either error in the flux estimates or a real net loss of nutrients within the 

subbasin (Aulenbach et al., 2007).

 To further examine source areas of N, P and silicate, the nutrient fluxes in the 

MARB were divided into ten smaller subbasins (see Figure 24 and Table 4), with some of 

the values calculated as the difference between an upstream and downstream monitoring 

station. The lower Mississippi River subbasin is again calculated by difference and is the 

same in both the five and ten subbasin analysis (this subbasin is not shown in Figure 1, 

but was included in the Table 3 analysis). These results are listed in Table 4. For nitrateN, this further breakdown of the basin indicates that the largest sources are the upper 
76

Mississippi and Ohio-Tennessee River subbasins. These subbasins represent about 31% 

of the total land area within the MARB, yet they contribute about 82% of the nitrate-N 

flux, 69% of the total Kjeldahl N, and 58% of the total P flux. Furthermore, when the 

subbasins are further divided, the subbasin contributing to the upper Mississippi River 

between Clinton, IA and Grafton, IL contributes about 29% of the nitrate-N flux, while 

representing only 7% of the drainage area. The Missouri River at Hermann also was a 

relatively large contributor of total P (14% of total flux). For dissolved silicate, 

percentages did not include the Red River because estimates were not available. Again, 

most of the silicate flux was from the upper Mississippi River and the Ohio-Tennessee 

River, similar in proportion to water flux. 

Table 4: Average annual nutrient fluxes for ten subbasins in the MARB for the 2001-2005 water years.. 

Some subbasin fluxes are calculated as the difference between the upstream and downstream monitoring 

station. (Percent of total basin flux shown in parentheses.)

Subbasin Area Nitrate-N TKN Total P Si 

(km2

) 

Flow 

(M m3/yr) 1,000 metric tons 

Mississippi-Clinton 222,000 48,300 88.3 (11%) 50.1 (12%) 8.5 (6%) 219 (12%) 

Mississippi-Grafton1 221,700 52,100 237 (29%) 71.7 (17%) 21.2 (14%) 162 (9%) 

Missouri-Omaha 836,000 23,900 24.1 (3%) 25.4 (5.9%) 8.1 (5%) 102 (6%) 

Missouri-Hermann1 517,000 36,100 54.6 (7%) 58.4 (14%) 22.3 (14%) 161 (9%) 

Mississippi-Thebes1 50,300 15,800 23.8 (3%) 13.9 (3%) 10.8 (7%) 8.5 (0.5%) 

Ohio-Cannelton 251,000 133,400 160 (20%) 92.1 (21%) 35.2 (23%) 355 (20%) 

Ohio-Grand Chain1 275,000 146,400 175 (22%) 82.7 (19%) 23.5 (15%) 320 (18%) 

Arkansas-Little 

Rock 

409,300 33,900 21.9 (3%) 19.5 (5%) 4.4 (3%) 102 (6%) 

Red RiverAlexandra 175,000 33,200 6.8 (1%) 24.3 (6%) 4.3 (3%) 

Lower Mississippi1 183,200 129,550 22.1 (2.7) -8.4 (-2) 16.1 (10%) 

757 (20%)2

1

For these basins, fluxes were calculated as the difference between upstream and downstream stations. 2

For these two subbasins, calculated by difference from overall basin flux minus eight subbasins where Si 

flux was estimated. 

 Annual Yield Estimates: Similarly, the nitrate-N and TKN yields were dominated 

by the Upper Mississippi and Ohio-Tennessee River subbasins, with nitrate-N values of 

7.1 and 6.4 kg N/ha/yr (6.3 and 5.7 lb N/ac/yr) and TKN values of 2.7 and 3.3 kg N/ha/yr 

(2.4 and 2.9 lb N/ac/yr) for the upper Mississippi and Ohio-Tennessee River subbasins, 

respectively (Table 5). The Missouri and Arkansas-Red River subbasins had much lower 

nitrate-N yields of 0.6 and 0.5 kg N/ha/yr (0.53 and 0.44 lb N/ac/yr)for this five-year 

period. Similar to N, yield of total P was much greater in the upper Mississippi and 

Ohio-Tennessee River subbasins when compared to the Missouri River. The greater 

yields from the upper Mississippi and Ohio-Tennessee River basins no doubt reflect the 

relative sizes of the basins when compared to the Missouri River but also the importance 

of point sources in the basins, as well as more intensive agricultural inputs. 
77

Table 5: Average annual nutrient yields for the five large subbasins in the MARB for water years 2001-

2005. 

Subbasin Nitrate-N TKN Total P 

(kg/ha/yr) 

Upper Mississippi 7.1 2.7 0.8 

Ohio-Tennessee 6.4 3.3 1.1 

Missouri 0.6 0.6 0.2 

Arkansas-Red 0.5 0.8 0.1 

Lower Mississippi 1.2 -0.5 0.9 

 When nutrient yields from the nine smaller subbasins are examined, the yields 

from the upper Mississippi River between Clinton and Grafton and the entire Ohio River 

basin were 10.7 and 6.4 kg N/ha/yr (9.6 and 5.7 lb N/ac/yr), respectively (Table 6). The 

largest total P yield (2.1 kg P/ha/yr or 1.9 lb P/ac/yr) was from the subbasin measured on 

the Mississippi River at Thebes, which would include row crop lands of Missouri River 

and southern Illinois River along with sewage effluent from St. Louis. Greatest dissolved 

silicate yields were from the Ohio River, followed by the upper and lower Mississippi 

River, again reflecting water flux. 

Table 6: Average annual nutrient yields for nine subbasins in the MARB for the 2001 - 2005 water years. 

Some subbasin yields are calculated as the difference between the upstream and downstream monitoring 

stations. 

Subbasin Nitrate-N TKN Total P Silicate 

(kg/ha/yr) 

Mississippi-Clinton 4.0 2.3 0.4 9.9 

Mississippi-Grafton 10.7 3.2 1.0 7.3 

Missouri-Omaha 0.3 0.3 0.1 1.2 

Missouri-Hermann 1.1 1.1 0.4 3.1 

Mississippi-Thebes 4.7 2.8 2.1 1.7 

Ohio-Cannelton 6.4 3.7 1.4 14.1 

Ohio-Grand Chain 6.4 3.0 0.9 11.6 

Arkansas-Little Rock 0.5 0.5 0.1 2.5 

Red River-Alexandra 0.4 1.4 0.2 

Lower Mississippi 1.2 -0.5 0.9 9.91

1

Flux calculation available only for sum of two subbasins. 

 Subbasin Nitrate-N Yield Compared to Net N Inputs: The complete time series 

records were examined to better understand longer term patterns in subbasins 

contributing the largest N and P fluxes. At the five subbasin level, the trend lines for 

flow and N fluxes for the Ohio River basin have been relatively flat since the early 1980s 

(Figure 25). 
78

Water Flux

(million m

3

)

0

100000

200000

300000

400000

500000

1940 1950 1960 1970 1980 1990 2000 2010

N Flux or Yield

(kg N ha-1 yr-1)

0

10

20

30

40

N Flux or Yield

(million tons N)

0.0

0.5

1.0

1.5

2.0

Net N Inputs

River Nitrate-N yield

Ohio River subbasin

Figure 25: Net N inputs and annual nitrate-N fluxes and yields for the Ohio River subbasin. (LOWESS 

curves for riverine nitrate-N shown in red.) Based on USGS data from Battaglin (2006) and Aulenbach et 

al. (2007).

However, the upper Mississippi River subbasin has experienced a decreasing trend in 

annual flow since the mid 1990s (Figure 26). What appears to be only a slight decrease 

in nitrate-N yield in the upper Mississippi subbasin in response to what the panel thinks 

are greatly decreasing net N inputs, demonstrates the difficulty in predicting riverine 

nutrient yields in tile-drained agricultural lands. Many interacting factors are at work, 

which are difficult to estimate and/or measure. For example, there are uncertainties in 

some of the estimates, such as biological N2 fixation (primarily soybean), as well as our 

assumption that large soil N pools are in a steady state. The predominant soil types in the 

upper Mississippi subbasin are Mollisols, which are high in organic matter with large soil 

organic N pools (much larger than the Ohio River subbasin). As fertilizer rates have 

stayed constant and yields have increased, several possibilities may account for the lack 

of riverine response. These include increasing soybean N2 fixation percentages, net N 
79

mineralization of soil organic N (David et al., 2001), long lag times due to a buildup of 

relatively easily degradable organic N (amino sugar N, Mulvaney et al., 2001) that is now 

being released, or perhaps increasing tile drainage and loss of fall applied N. Figure 26 

includes a recalculation of net N inputs for 1998 to 2005, increasing soybean fixation 

rates from 50 to 70%, and assuming a corn acre net soil mineralization rate of 10 kg 

N/ha/yr (8.9 lb N/ac/yr). These two changes greatly alter the net inputs, pushing the 

value back up to where it was during the 1980s. Water Flux (million m

3

)

0

50000

100000

150000

200000

250000

Upper Mississippi River subbasin

1940 1950 1960 1970 1980 1990 2000 2010

N Flux or Yield

(kg N ha-1 yr-1)

0

10

20

30

40

N Flux or Yield

(million tons N)

0.0

0.5

1.0

1.5

2.0

Net N Inputs

Riverine Nitrate-N yield

Net N Inputs revised

Figure 26: Net N inputs and annual nitrate-N fluxes and yields for the upper Mississippi River subbasin. 

(LOWESS curves for riverine nitrate-N shown in red.) Shown in green is a recalculated net N input for the 

upper Mississippi River basin, increasing soybean N2 fixation from 50 to 70% of above ground N, and a 

soil net N mineralization rate from 0 to 10 kg N/ha/yr. Based on USGS data from Battaglin (2006) and 

Aulenbach et al. (2007). 

 Soybean production is a net depletion to soil N pools and the fixation rate is a 

function of available inorganic N (nitrate) in the soil (Gentry et al., 2001). When there 

was more inorganic N left from corn production prior to the late 1990s, soybeans would 

have fixed less N compared to recent growing seasons when corn yields have set records, 

and little residual soil nitrate would be expected. This could be leading to increasing 

soybean N2 fixation rates, which are not accounted for in typical net N input calculations. 
80

 A second factor is soil mineralization. Net N input calculations assume that the 

soil organic N pool is at a steady state (McIsaac et al., 2002), with mineralization rates in 

a year balanced by immobilization (both microbial and crop residue inputs). It is possible 

that with greater corn production and steady fertilizer rates, increased mineralization rates 

occur, so that there is a net depletion of soil organic N (one component of soil organic 

matter, which is discussed further in Section 4.5.6). This depletion, as discussed earlier, 

may be small (about 10 kg N/ha/yr or 8.9 lb N/ac/yr) but over many acres would be an 

important additional input. 

 Finally, another factor may be an increase in tile drainage intensity in the region, 

combined with increasing fall fertilization and warmer winter temperatures. New and 

replacement tile drainage is added every year to this region, although no data are 

available to quantify the increase. Fall application of anhydrous ammonia in much of the 

region has increased greatly since the 1980s (see later discussion in Section 4.5.6 for 

supporting sales and USDA ARMS data). The four states of the upper Mississippi River 

basin (Minnesota, Wisconsin, Iowa and Illinois) all show an increasing winter 

(November through March) temperature (for the months following fall application of 

anhydrous ammonia all show strong increasing trends in winter temperatures during the 

last 30 years, data not shown). Warmer soils would increase nitrification rates and lead 

to higher concentrations of soil nitrate that could be lost with late winter and spring 

precipitation. Therefore, fall applied anhydrous ammonia could be a more important 

source of spring nitrate-N flux in this subbasin during recent years and, when combined 

with changing N input and output patterns, may be keeping the flux steady despite the 

reduction in annual net N inputs. 

 Changes in subbasin P: As discussed previously, total P flux for the MARB has 

increased during the monitoring period. Most of this increase was found to have 

occurred in the Ohio River subbasin, particularly during the 2001 to 2005 time period 

(Figure 27). In comparing the 2001 to 2005 period with 1980 to 1996, Ohio River total P 

increased 51%, while water flux increased only 6%, and reactive P decreased by 20%. 

This led to a large increase in particulate/organic P of 89% between these two time 

periods. Because TKN decreased by 3% during this period, it does not seem that 

increased erosion can explain this pattern (all indications are that erosion has decreased). 

The 89% increase in particulate/organic P represents most of the increase in total P flux 

to the NGOM between 1980 to 1996 and 2001 to 2005. Unfortunately, data are not 

available because of monitoring limitations for smaller basins within the Ohio River 

subbasin to further determine the source of this P flux. However, this trend seems to be 

more widespread than just the Ohio subbasin. 
81

1975 1980 1985 1990 1995 2000 2005 2010

Riverine P Flux

(metric tons P yr-1)

0

20000

40000

60000

80000

100000

Total P 

Particulate/organic P 

Figure 27: Total P and particulate/organic P fluxes for the Ohio River near Grand Chain, Illinois. 

(LOWESS curves shown in black and red). Based on USGS data from Battaglin (2006) and Aulenbach et 

al. (2007).

The Missouri and Upper Mississippi River subbasins are following a similar trend 

as the Ohio River, although their absolute increase in total P is much less than the Ohio 

River. In both of these subbasins flow has decreased (by 10 and 31% for the Upper 

Mississippi and Missouri River subbasins, respectively for 1980 to 1996 compared to 

2001 to 2005), while total P flux has increased (about 10% in each subbasin). Again, 

TKN flux has decreased. Therefore, in the Missouri, Upper Mississippi, and Ohio River 

subbasins flow weighted total P concentrations have increased greatly during the last 15 

years. 

 These observations are not consistent with overall TKN riverine fluxes in the 

MARB, and at this time the SAB Panel has no explanation for this large, yet potentially 

very important, change in total P concentrations and flux for these subbasins which could 

influence management decisions. 

Seasonal Patterns 

 Spring fluxes (sum of April, May, and June) were examined for the Mississippi 

River at Grafton and the Ohio River at Grand Chain, and little change in water flux was 
82

detected (Figure 28). However, for nitrate-N, there seems to be a slight increasing 

pattern of spring flux based on LOWESS curves. Water Flux (million m3

)

0

20000

40000

60000

80000

0

40000

80000

120000

160000

Mississippi River at Grafton

Ohio River at Grand Chain, IL

Riverine Nitrate-N Flux 

(million metric tons N)

0.00

0.05

0.10

0.15

0.20

0.25

0.30

1975 1980 1985 1990 1995 2000 2005

0.00

0.05

0.10

0.15

0.20

0.25

0.30

Mississippi River at Grafton

Ohio River at Grand Chain, IL

Figure 28: Spring water flux and nitrate-N flux for the Mississippi River at Grafton and the Ohio River at 

Grand Chain, IL for water years 1975-2005. (LOWESS curves shown in red.) Based on USGS data from 

Battaglin (2006) and Aulenbach et al. (2007). 

 When the sum of the upper Mississippi River at Grafton and Ohio River at Grand 

Chain spring nitrate-N flux is plotted against the flux for the entire basin, an interesting 

pattern emerges (Figure 29). During the 1980s into the early 1990s, some of the spring 

flux was from other subbasins, mostly the Missouri River. However, the Missouri River 

flux has greatly decreased so that now the upper Mississippi River above Grafton and the 

Ohio River contribute nearly all of the spring flux. Sprague et al. (2006) discuss the 
83

riverine fluxes in the Missouri River basin (due to decreasing flow and management 

practice changes) in a recent report that supports this observation. 

1975 1980 1985 1990 1995 2000 2005

Riverine Nitrate-N Flux 

(million metric tons N)

0.0

0.2

0.4

0.6

0.8

1.0

MARB Basin

Grafton + Ohio

Figure 29: Spring nitrate-N flux (sum of April, May, and June) for the Mississippi River at Grafton plus 

Ohio River at Grand Chain subbasins compared to the combined Mississippi and Atchafalaya River for 

1979 through 2005. Based on USGS data from Battaglin (2006) and Aulenbach et al. (2007). 

3.1.3. Key Findings and Recommendations on Temporal Characteristics 

 Most of the research needs identified in the Integrated Assessment have not been 

met, and fewer rivers and streams are monitored today than in 2000. Data continue to be 

available for the large river sites, but many intermediate and smaller river monitoring 

sites have been dropped from monitoring programs. Recently USGS has initiated real 

time (every two hour) monitoring of three large river sites with field nitrate-N 

measurement. These types of new efforts to provide expanded monitoring data are 

critically needed. To more fully assess the response of the entire suite of management 

programs and changes at the subbasin and large river scale in the MARB, we need more 

robust monitoring programs that have adequate sampling intensities to allow the 

composite method (the preferred one) of estimating stream loads to be utilized. At the 

small watershed (1,000 to 50,000 ha or about 2,500 to 125,000 ac) scale, there have been 

many studies, but they provide data for only the period of funding, which is often short. 

A monitoring network is needed throughout the MARB focused on small watersheds with 

larger N and P loads and that provides intensive, long-term data. This network will allow 

determination of how effective particular individual or suites of management programs 

are in reducing nutrient loads. However, because of year-to-year weather patterns and 

the often slow response of changes in outputs, these programs will need to be in place for 

decades. Finally, there is a critical need for the ability to document tile drainage 

intensity, which requires that new techniques be developed and applied. 
84

 Changes in USGS flux calculation methods have altered estimates of nutrient flux 

as reported in the Integrated Assessment. LOADEST 5 yr and a new COMPOSITE 

method seem to be the best estimation methods. Although water flux for the MARB has 

increased slightly during the past 25 years, total N, primarily nitrate-N and 

particulate/organic N, has decreased. The total N flux averaged 1.24 million metric 

tons/yr (1.37 million tons/yr) from 2001 – 2005 (65% of the flux is nitrate), and the total 

P flux averaged 154,000 metric tons/yr (170,000 tons/yr). During the spring (AprilJune), water flux for the MARB appears to have decreased slightly, causing similar 

decreases in total N (nitrate-N and TKN). Spring dissolved silicate flux has declined 

more than water flux. Neither total P nor SRP fluxes show major annual or seasonal 

trends during the full period of record. 

 The subbasin analysis provides clear evidence that while the upper Mississippi 

and Ohio-Tennessee River subbasins represent about 31% of the total drainage area of 

the MARB, they contribute about 82% of the nitrate-N flux, 69% of the TKN flux, and 

58% of the total P flux to the Gulf. Furthermore, when the subbasins are further divided, 

the subbasin contributing to the upper Mississippi River between Clinton, IA and 

Grafton, IL contributes about 29% of the nitrate-N flux while representing only 7% of the 

drainage area. Perhaps more importantly, the upper Mississippi and Ohio-Tennessee 

River subbasins currently represent nearly all of the spring N flux to the Gulf. These 

subbasins represent the tile-drained, corn-soybean landscape of Iowa, Illinois, Indiana, 

and Ohio and illustrate that corn-soybean agriculture with tile drainage leaks considerable 

N under the current management system. The source of riverine P is more diffuse, 

although these subbasins are also the largest sources of P. A large increase in the Ohio 

River subbasin particulate/organic P flux occurred during the 2001 to 2005 time period, 

which was the source of nearly all of the increase in total P to the NGOM. At the same 

time flow weighted total P concentrations increased in the Upper Mississippi and 

Missouri River subbasins as well, although increases in flux were smaller than the Ohio 

River due to decreased water flux. The SAB Panel has no explanation for this striking 

change in P concentrations in these subbasins. 

 Based on these findings, the Panel recommends the following: 

• Establishment of a monitoring network (20 to 100) of small watersheds will 

provide long-term (tens of years), intensive flux data to determine the response of 

management programs and decisions in the MARB. 

• More intensive monitoring of larger rivers at the subbasin and entire MARB scale 

is needed to allow for monthly calculation of fluxes using the composite 

estimation method, the most accurate method estimating fluxes. 

• Further research is needed to determine why riverine spring nitrate-N fluxes are 

not declining in response to annual net N input decreases, which will inform 

management decisions for corn/soybean agriculture. 
85

• The increase in riverine total P concentrations needs to be fully explored to verify 

the increase and to further document the source, potentially having great 

management implications for control of P in the MARB. 

• The tile-drained Corn Belt region of the MARB is an important target for 

reductions in both N and P, focusing on both surface (P) and sub-surface losses 

(N). 

• Additional research is needed to better define the extent, pattern, and intensity of 

agricultural drainage, including cropland drained by field tile as well as cropland 

not directly drained by field tile but contributing to drainage networks. 

3.2. Mass Balance of Nutrients 

 Mass balance can be used to better understand sources, sinks, and transformations 

of nutrients in ecosystems, although losses to stream water are not specifically 

determined. Goolsby et al. (1999) constructed a detailed annual N mass balance for 1960 

– 1996 and a P mass balance for 1992. Improving flux estimates was identified as a 

research need. In particular, better estimates are needed for soil N mineralization, soil 

immobilization, plant N volatilization, denitrification, and biological N2 fixation. 

Cropping Patterns 

 Mass balances reflect the types and areas of crops grown across the MARB. 

There were large changes in these crops over the past half century (Figure 30). Earlier 

cropping systems had more diverse rotations, including corn, wheat, hay, and oats. With 

the onset of modern agriculture and large fertilizer inputs, much of the MARB is now in a 

corn and soybean rotation. By the late 1990s, corn and soybean areas were equal but 

more recently corn acreage has increased and soybean has decreased, with this trend very 

apparent in 2007. This trend is expected to continue as demand for corn increases due to 

expanding ethanol production, the implications of which are discussed in detail in Section 

4.5.9. 
86

1940 1950 1960 1970 1980 1990 2000 2010

Crop Area (1000 ha)

0

5000

10000

15000

20000

25000

30000

35000

Corn

Soybean

Hay

Oat

Wheat

Figure 30: Area of major crops planted in the MARB from 1941 through 2007. Adapted from McIsaac, 

2006. 

Non-Point Sources 

 Nitrogen: The N mass balance described in the Integrated Assessment indicated 

that there was a greater surplus of N during the 1950s than during the 1980s and 1990s 

(Goolsby et al., 1999). McIsaac et al. (2001, 2002) used the same data set to determine 

the N mass balance using a method described by Howarth et al. (1996) that also has been 

used by many others (e.g., David and Gentry, 2000; David et al., 2001; McIsaac and Hu, 

2004 for Illinois). Net anthropogenic N inputs (NANI) were calculated (sum of fertilizer, 

NOy deposition, N2 fixation, minus net food and feed imports) from existing MARB data 

bases, assuming that the large soil organic N pool is in a steady state. Manure is included 

in this calculation as part of the feed imports, where grain consumed and excreted as a 

part of animal agriculture is estimated. NANI is N that should be available for 

denitrification, loss to groundwater, or leaching and transport in streams. 

 The recalculated NANI for the MARB showed a clear increase from about 9 kg N 

ha/yr (8 lb N/ac/yr) in the 1940s to about 16 kg N/ha/yr (14 lb N/ac/yr) from the early 

1980s to present, with a maximum value of 20.9 kg N/ha/yr (18.7 lb N/ac/yr) in 1988 

(Figure 31). This increase was due to increasing fertilizer N inputs (from 0 to ~ 20 kg 

N/ha/yr or 17.9 lb N/ac/yr)) and higher N2 fixation from the increased soybean 

production (from about 8-14 kg N/ha/yr or about 7-12.5 lb N/ac/yr). Atmospheric 

deposition appears to be the greatest in the Ohio River basin (about 16% of NANI) and 

shows a slight increase basin-wide but generally is a small component of the NANI (for a 
87

more detailed discussion see Appendix B: Flow Diagrams and Mass Balance of 

Nutrients. Manure shows a slight decrease across the MARB, as extensive animal 

production has moved to feedlots further west, but represents only about 16% of the total 

inputs. However, animal production has become concentrated in specific regions of the 

MARB, creating localized nutrient surpluses compared with crop needs and offtake 

(USDA, 2003). Up to now, this has led to water quality impairment at a local rather than 

MARB scale, due to where the animal operations have become concentrated (for more 

information on distribution see Section 4.5.5 and Appendix E: Animal Production 

Systems). Therefore, the major changes in inputs were due to fertilizer and N2 fixation. 

However, when compared to the amount of N removed during crop harvest, which has 

dramatically increased since 1940, the increase in N inputs from fertilizer and N2 fixation 

don’t appear to have increased proportionately. In fact, this rapid increase in crop 

production has led to a small decrease in NANI from about 17 kg N/ha/yr (15 lb N/ac/yr) 
88

in 2000, to net N inputs of 14 kg N ha/yr (12.5 lb N/ac/yr) in 2004 and 2005 (McIsaac, 

0

10

20

30

40

Fertilizer 

N2 fixation 

NOy deposition 

Nitrogen Balance Component (kg N ha-1 yr-1)

0

10

20

30

40

Manure 

Human consumption

Harvest 

1940 1950 1960 1970 1980 1990 2000 2010

0

10

20

30

40

Annual value

Previous 2 to 5 year average

Net Nitrogen Inputs

Figure 31: Nitrogen mass balance components and net N inputs for the MARB, as calculated by McIsaac 

et al. (2002) and updated through 2005 by McIsaac (2006). 

2006). 

 The subbasins that contribute the greatest N flux to the Gulf are the upper 

Mississippi and Ohio River basins, due largely to the intensity of agriculture with 

concomitant large inputs of N from fertilizer and fixation combined with the system of 

tile drains. Therefore, when the nitrogen balance is presented by subbasin (Figure 32) the 

highest net nitrogen inputs are to those subbasins. 
89

0

10

20

30

40 Ohio 

Upper Miss

Net Nitrogen Inputs (kg N ha-1 yr-1)

0

10

20

30

40

Missouri 

Lower Miss 

1940 1950 1960 1970 1980 1990 2000 2010

0

10

20

30

40

Entire MRB

Figure 32: Net N inputs for the four major regions of the MARB through 2005. Adapted from McIsaac, 

2006. 

 However, a closer look at the inputs to the upper Mississippi River basin shows 

that, even though N inputs from fertilizer and N2 fixation appear to be fairly level during 

recent years, the amount of N removed during harvest continues to increase, resulting in a 

substantial decline in NANI (Figure 33). These changes are not reflected in the other 

subbasins, which lead to a small decline in NANI to the overall basin. However, given 

the importance of the upper basin as a source of nitrate-N, it might be expected that the 

riverine flux of N would start to decrease. 
90

0

10

20

30

40

50

Fertilizer

N2 Fixation

NOy 

Nitrogen Balance Component (kg N ha-1 yr-1)

0

20

40

60

80

Manure 

Human consumption

Harvest

1940 1950 1960 1970 1980 1990 2000 2010

0

10

20

30

40

50

Annual value

Previous 2 to 5 yr

 average

Net Nitrogen Inputs

Figure 33: Nitrogen mass balance components and net N inputs for the upper Mississippi River basin, as 

calculated by McIsaac et al. (2002) and updated through 2005 by McIsaac (2006).

 McIsaac et al. (2001, 2002) showed that net N inputs could be used, in 

combination with riverine water flux, to predict export of nitrate-N to the Gulf. They 

found that a 2-5 year lagged net N input explained the most variation in nitrate-N export, 

with 6-9 year lagged net N inputs explaining less, but a significant amount of the 

variation. Therefore, given the large decrease in net N inputs in the upper Mississippi 
91

River subbasin, it is reasonable to expect riverine export of nitrate should decrease. 

However, there is a factor that is not assessed in the net N input mass balance that may be 

important. 

 McIsaac and Hu (2004) showed that, for tile and non-tile drained regions of 

Illinois, net N inputs were similar but that riverine export of N was much greater in the 

tile drained watersheds. They found that during the 1990s net N inputs were equal to 

riverine N flux, about 27 kg N ha/yr (24 lb N/ha/yr). This would leave no N available for 

other fluxes that are thought to be important, such as terrestrial and aquatic 

denitrification. More recent net N inputs in these same tile drained watersheds are about 

zero, yet riverine N export has continued. Given that there are denitrification losses (that 

are unmeasured), this result indicates that N must be coming from a depletion of soil N 

pools, as suggested by Jaynes et al. (2001). With steady fertilizer N rates, high corn and 

soybean yields, and high stream N export, the only source available to supply N would be 

the large soil N pool (often 10,000 to 15,000 kg N/ha or 8,930 to 13,400 lb N/ac) in the 

Mollisols of the upper Midwest. Techniques are not yet available to document the small 

change that would be occurring in this N pool from a small annual depletion of 25 to 50 

kg N/ha/yr (22 to 45 lb N/ac/yr); however, this possibility has critical implications for the 

sustainability of production. 

 Another possibility raised by McIsaac et al. (2002) is that estimates of crop 

harvest N, N2 fixation, or animal consumption of N and manure production could be 

inaccurate. Although Goolsby et al. (1999) recommended that we improve estimates of 

the N mass balance, we have not made progress in our methods or data available to 

calculate individual fluxes of N. Manure is an important component of the mass balance 

and can be thought of as N that is not exported in grain (or forage that is consumed) or, 

therefore, the N that is returned to the landscape in the MARB. There are many 

assumptions in calculating the manure flux that could also alter our interpretation of the 

overall mass balance. 

 Phosphorus: A P mass balance for 1992 was included in the Integrated 

Assessment that incorporated fertilizer, manure, grain harvest, hay harvest, and pasture 

grazing (Goolsby et al., 1999). Small but potentially important changes in the large soil 

pool were not included because methods are not available for making this estimate for 

short-time spans. 

 A P mass balance was calculated using the extended N mass balance (McIsaac, 

2006) for 1951 - 2005 for each state, and these values were then summed for the MARB 

(Figure 34). P fertilizer inputs have decreased since the 1970s such that the increased 

harvest now exceeds fertilizer inputs (and manure retention) most years, so large soil P 

pools are being utilized by crops. The large buildup of soil P in the 1970s and 1980s led 

to a large positive net P balance, but decreased fertilizer inputs and high crop yields result 

in the current negative balance. 
92

0

2

4

6

8

10

Phosphorus Balance Component (kg P ha-1 yr-1)

0

2

4

6

8

10

Manure 

Human Consumption

Harvest 

1950 1960 1970 1980 1990 2000 2010

-1

0

1

2

3

Net P Inputs

Fertilizer

Figure 34: Phosphorus mass balance components and net P inputs for the MARB. Adapted from McIsaac, 

2006. 

 When P mass balance is calculated for major subbasins, only the Lower MARB 

still has a positive P balance (Figure 35). The Missouri River P balance has shown little 

change, while the Ohio and Upper Mississippi River have a negative P balance. In 

contrast to N, the amount of P lost to streams and exported by rivers is small relative to 

agronomic fluxes; hence it is not expected that these changes in P mass balances will 

cause short-term (or even relatively long-term) changes in stream P concentrations and 

loads (David and Gentry, 2000). 
93

-2

0

2

4

6

Ohio 

Upper Miss

Net Phosphorus Inputs (kg P ha-1 yr-1) -2

0

2

4

6

Missouri 

Lower Miss 

1950 1960 1970 1980 1990 2000 2010

-2

0

2

4

6

Entire MRB

Figure 35: Net P inputs for the four major subbasins of the MARB through 2005. Adaptive from McIsaac, 

2006. 
94

 A closer look at the upper Mississippi River basin (Figure 36) shows an even 

larger decline in P from fertilizer and a steady decline in P from manure. 

0

5

10

15

Phosphorus Balance Component (kg P ha-1 yr-1)

0

5

10

15

Harvest

Manure

Human

1950 1960 1970 1980 1990 2000 2010

-2

0

2

4

6

Net P Inputs

Fertilizer

Figure 36: Phosphorus mass balance components and net N inputs for the upper Mississippi River basin. 

Adapted from McIsaac, 2006. 
95

Point Sources 

 In the Integrated Assessment, point sources were estimated to contribute about 

11% of the total nitrogen and an undefined, though likely somewhat lower, total 

phosphorus flux to the MARB. This assessment (Tetra Tech, Inc., 1998) was based on 

1996 information, and it estimated fluxes at 321,000 metric tons N/yr (354,000 tons N/yr) 

and 91,500 metric tons P/yr (101,000 tons P/yr). 

 A reassessment (MART, 2006b) was based on 2004 permit information, adjusted 

assumptions, evaluated more facilities, and revised estimated fluxes downward to 

233,000 metric tons N/yr (257,000 tons N/yr) and 39,500 metric tons P/yr (43,500 tons 

P/yr). Municipal treatment plants (STP) were thought to account for about 65% of the 

total point-source fluxes for both N and P. However, few permits have suitable data for 

direct flux calculations, and only 11.1% of the mass flux was directly calculated from the 

permit information. The rest of the mass flux was estimated using “typical pollutant 

concentrations” (TPC) and estimated daily water flows from point sources. The TPCs 

used in the MART (2006) estimates are lower than those used by other water quality 

programs, therefore, the SAB Panel has re-calculated the contribution of N and P from 

municipal sewage treatment plants based on effluent concentrations that better reflect 

measured nutrient concentrations from point sources during 2004. These calculations 

also assume that the point source load is delivered to the NGOM without any in-stream 

losses. Therefore, they are the upper estimate for the contribution of point sources to the 

total N and total P riverine load. The SAB Panel’s calculation indicates that load 

estimates would need to be revised upward to 267,000 metric tons N/yr (294,000 tons 

N/yr) (72% from STPs and 28% from industrial sources) and 53,000 metric tons P/yr 

(58,500 tons P/yr) (77% from STPs and 23% from industrial sources). (See Appendix D 

for a more detailed discussion of the SAB Panel’s estimates.) When the contributions 

from all point sources are compared to the average annual N and P fluxes for the period 

2001 – 2005, these new estimates indicate that point sources contribute to the Gulf about 

22% and 34% of the average annual N and P flux, respectively. When compared to 2004 

N and P fluxes (slightly higher than average fluxes), the percentage of the N flux 

contributed by point sources drops to about 20%, and the P flux remains constant at about 

34%. Fluxes from point sources are equally distributed throughout the year, but spring 

flux is critical to the Gulf. Assuming equal monthly loads from point sources, the SAB 

Panel’s estimates indicate that point sources are responsible for approximately14% of 

spring N flux and 27% of spring P flux for 2001 - 2005. Again, the Panel emphasizes 

that these are rough estimates, as measured data are not available at this time to make 

more accurate determination of point source contributions. 

 A summary of the percent of P fluxes by major hydrologic region, based on the 

new estimates, is shown in Figure 37. Collectively, the upper Mississippi and Ohio River 

basins account for about half the P flux from point sources in the MARB. 
96

Unknown

2%

05 Ohio

24%

06 Tennessee

5%

07 Upper 

Mississippi

27%

08 Lower 

Mississippi

13%

10 Missouri

16%

11 ArkansasRed-White

13%

Figure 37: Total phosphorus point source fluxes as a percent of total flux for the MARB for 2004 by 

hydrologic region. 

 This analysis suggests that point source P fluxes are a significant source of both 

annual and spring fluxes to the MARB and the Gulf and that substantial reductions in P 

fluxes in the MARB are likely if P fluxes from point sources are reduced. Point sources 

are a less important source of spring and annual N flux; however, reduction in N fluxes 

from point sources may offer a certain and cost effective means of achieving some of the 

N reductions needed in the MARB. It is important to emphasize that the differences in 

assumptions used to estimate fluxes based on TPC have a major impact on annual and 

seasonal flux estimates for the MARB and would likely affect the estimated cost 

effectiveness of requiring N or P removal from point sources in the MARB (discussed 

further in Section 4.5.8). 
97

Key Findings and Recommendations 

 Although N mass balances have been recalculated since the Integrated 

Assessment, the research needs described in that report remain. Components of the N 

mass balance such as denitrification, N2 fixation, manure N, and soil N pool processes 

such as mineralization and immobilization are not measured each year. Only N2 fixation 

and manure N can even be estimated, with the other fluxes having little data available to 

make calculations. Point sources export N and P directly to rivers, yet their contributions 

continue to be estimated from permits. 

 New methods have been used to calculate N mass balances in this report (net 

anthropogenic N inputs, NANI). NANI and net P inputs for MARB have increased 

greatly since the 1950s; but have decreased in the past decade because of steady fertilizer 

applications and increased crop yields for N, and reduced fertilizer applications and 

increased crop yields for P. Mass balances in the upper Mississippi River subbasin 

suggest that under the current tile drained corn and soybean management system 

depletion of soil organic N pools may be occurring. From a sustainability viewpoint, this 

needs to be fully documented and decreased as new systems are put in place to reduce N 

export in rivers. Point sources represented 22% of riverine N flux and 34% of P flux 

delivered to the Gulf. Manure is a more significant source of P than N; and where 

riverine N flux is greatest, excess manure N tends to be a less important input. Manure is 

likely more important basin-wide to local water quality problems, rather than a large 

component of MARB export of N or P, because of where concentrated animal production 

has relocated. The greatest decrease in net N and P inputs was seen in the upper 

Mississippi River basin. From 1999 - 2005, 54% of N inputs were from fertilizer, 37% 

from fixation, and 9% from deposition for the entire basin. Deposition was most 

important in Ohio basin (16% of inputs). Based on these findings, the Panel offers these 

recommendations. 

• Continue and expand research to more accurately and fully measure the N 

mass balance in the MARB by developing methods and gathering data for 

improving the estimates of critical fluxes such as N2 fixation, manure, 

denitrification, and soil N pool changes. 

• Sustainability of soils in the MARB must be fully addressed by research to 

improve measurement of changes in soil N pools as a result of new 

management systems, with changes in soil N pools incorporated into more 

complete N mass balances. Section 4 discusses the need for research on 

changes in N pools associated with different management practices, e.g. 

tillage systems and other practices. 

• N and P from point sources should be estimated from direct measurements, 

rather than relying on estimated values based on permits, so that more 

accurate calculations can be made of their contributions to the nutrient fluxes. 
98

3.3. Nutrient Transport Processes 

Aquatic processes 

 Studies conducted since the Integrated Assessment have addressed many of the 

research needs that were identified for nutrient transport processes: quantification of instream processes such as denitrification (particularly in small streams), research in small 

watersheds to identify dynamics and timing of N transport and to better understand the 

impact of drainage practices on nutrient flux, and development of a better understanding 

of N behavior during floods. We review these advances for nitrogen, phosphorus, and 

silicate transport and transformation. 

 Nitrogen. In-stream nitrogen removal in river networks is variable, but it can be 

substantial, particularly in river networks with relatively low nitrogen concentrations. In 

sixteen river networks in the northeastern United States, the Riv-N model predicted that 

37 to 76% of nitrogen inputs were removed within streams (Seitzinger et al., 2002), and 

the SPARROW model predicted that 7 to 54% of nitrogen inputs were removed 

(Alexander et al., 2002b). Estimates of the percentage of annual N inputs removed by instream processes in regional drainages in the Mississippi River basin range from 20 to 

55% (SPARROW model, Figure 38). The Ohio and White River basins removed the 

lowest percentage and the Arkansas and Missouri River basins the highest. Although 

these are estimates of the role of in-stream processes on an annual basis, the SPARROW 

model results strongly reflect the effects of seasonal pulses, especially the high spring 

values, because the mean annual flux is a flow-weighted estimate (R. Alexander, personal 

communication). 

Figure 38: Percentage of nutrient inputs to streams that are removed by instream and reservoir processes as 

predicted by the SPARROW model (Alexander et al., in press). 
99

 In-stream N removal accounts for a much smaller fraction of annual N export in 

tile-drained agricultural regions and other areas where stream water nitrogen 

concentrations are extremely high and water residence time is short. The proportion of 

the nitrate flux that was denitrified was highest in forested systems, lowest in urban, and 

intermediate in agricultural streams in Michigan (Inwood et al., 2005). Denitrification 

removed a greater fraction of N in meandering than in channelized reaches, but removal 

never exceeded 15%/day except during periods of low flow and warm temperature 

(Opdyke et al., 2006). Denitrification is a significant pathway for N removal in midwestern tile-drained streams during low flow, warm periods (summer and autumn), which 

improved local water quality at those times (Royer et al., 2004; Schaller et al., 2004). 

However, most of the nitrate is exported to the Gulf during high flows from January to 

June (Royer et al., 2006), and denitrification removes an insignificant fraction of this flux 

(Royer et al., 2004; 2006). Because in-stream removal is a small fraction of total flux at 

high flows, enhancing N removal by 50% during low flows (Q < median) would reduce 

annual N export only by less than 2% in Illinois agricultural streams; whereas enhancing 

removal by 25% during high flows (greater than 75th percentile flows) would reduce 

annual N export by 21% (Royer et al., 2006). 

 Recent research on streams in predominantly forested watersheds has shown that, 

in comparison to larger rivers, small streams remove a higher proportion of their 

incoming nitrogen per unit of water travel time (Alexander et al., 2000), per stream reach 

(Seitzinger et al., 2002), and per unit length (Wollheim et al., 2006; Helton, 2006). 

However, larger streams remove larger masses of nitrogen because more nitrogen passes 

through them (Seitzinger et al., 2002, Wollheim et al., 2006, Helton, 2006). Small 

streams receive and transport a significant amount of N to larger rivers, e.g. N loads to 

headwaters account for 45% of the load delivered to the entire river network in the 

northeastern U.S. (Alexander et al., 2007a). Similar calculations have not yet been done 

for the Mississippi River basin (R. Alexander, personal communication). Enhancing 

nutrient removal in small streams by restoring stream length that has been lost to 

straightening or burial could improve local water quality and decrease both N and P load 

to larger rivers (Bernot and Dodds 2005); however these reductions would be greatest at 

low stream flows, and less effective at high discharges when the bulk of nutrient load is 

being transported to the Gulf. 

 Denitrification is not the only pathway for N removal in streams, although it is the 

most permanent. Removal of nitrate from stream water and its assimilation into 

biological tissues transforms N from dissolved to particulate form, which reduces the rate 

at which it is transported downstream. Particulate N can be deposited and stored in 

sediments, where it can be mineralized and potentially denitrified. 

 Effectiveness of N removal in aquatic systems increases with water residence 

time, so reservoirs can make a significant contribution to N removal in river networks. 

Denitrification in an Illinois reservoir reduced average annual N export by 58%, but the 

percent reduction in annual export over a 23-year period varied from 31 to 91 % as 

retention time increased (David et al., 2006). N retention in Illinois reservoirs is higher 
100

than observed from rivers and reservoirs with lower nitrate concentrations (Figure 39). 

The difference can be attributed to lower removal efficiencies in natural lakes than in 

reservoirs where elevated inputs of N support high rates of denitrification in the 

sediments (David et al., 2006). Denitrification in aquatic sediments (80% in reservoirs in 

the tile-drained part of Illinois and 20% in streams) was estimated to reduce N export 

from Illinois by 25 % (David et al., 2006). Existing floodplain backwaters on the upper 

Mississippi River basin are limited in their effectiveness in N removal by denitrification 

because of short water-retention times and a lack of hydrologic connectivity with the 

main stem (Richardson et al., 2004; David et al., 2006). Enhancing connectivity and 

water-residence time on floodplains during periods of high discharge and high nitrate 

concentrations in the spring has been suggested as an effective way to reduce N loading 

to the Gulf (David et al., 2006). 

Depth/Time of Travel (m/yr)

0.1 1 10 100 1000 10000 100000

N Removed (%)

0

20

40

60

80

100

Lake Shelbyville

Seitzinger et al. rivers

Garnier et al.

Opdyke et al.

Royer et al.

Seitzinger Eq. 2

Our equation

y=88.453x-0.3677

r

2

=0.73

y=230.2.0x-0.5440

r

2

=0.91

Figure 39: N removed in aquatic ecosystems (as a % of inputs) as a function of ecosystem depth/water 

travel time (modified from David et al., 2006). Values shown are for 23 years in an Illinois reservoir (David 

et al., 2006), French reservoirs (Garnier et al., 1999), Illinois streams (an average from Royer et al., 2004), 

agricultural streams (Opdyke et al., 2006), and rivers (Seitzinger et al., 2002). The curve from Seitzinger et 

al. (2002) is not as steep as the curve that includes information from reservoirs in an agricultural region. 

 Because N2O is a potent greenhouse gas, whether the end product of 

denitrification is N2O or N2 is of importance. The IPCC estimates that 1.25% and 0.75% 

of N that enters agricultural soils and rivers, respectively, is converted to N2O (Mosier et 
101

al., 1998). However, that fraction includes N2O production via both nitrification and 

denitrification. IPCC assumes that only 0.5% of N that is denitrified in rivers is 

converted to N2O (Mosier et al., 1998), but they do not estimate this fraction for soils. A 

review of 32 studies of terrestrial denitrification reported the fraction of denitrified N 

converted to N2O to be highly variable (0 to 100%) with a mean of 27% (Stevens and 

Laughlin, 1998). Thus available data suggest that denitrification in aquatic systems 

produces less N2O as a fraction of denitrified N than terrestrial systems. Therefore, 

where denitrification occurs on the landscape will influence its contribution to 

greenhouse gases. However, enhancing denitrification to reduce water quality impacts of 

leached nitrogen will increase greenhouse gas emissions if nitrogen leaching rates remain 

high. 

Phosphorus. An understanding of P transport and transformation in streams and 

rivers has developed in parallel with the studies on N just described. Stream networks 

alter the timing, magnitude, and bioavailability of edge-of-field P loss during transport to 

the Gulf via geochemical and biological processes: sediment sorption and desorption, 

precipitation and dissolution, microbial and algal uptake, and riparian floodplain and 

wetland retention. Many of the geochemical processes are mediated by biota; e.g., coprecipitation of dissolved P with calcite may be biologically mediated during active 

photosynthesis (Neal, 2001), and aquatic biota accounted for 30 - 40% of sediment P 

uptake and release in wetland (Khoshmanesh et al., 1999) and stream sediments 

(McDowell and Sharpley, 2003). 

 Fluvial sediments come from overland flow and erosion of stream channels and 

banks. High discharge events that generate overland flow in agricultural regions 

commonly account for most of the annual phosphorus load (e.g., Gentry et al., 2007). 

Soils eroding from stream banks may be subsoils poor in P, which is less available for 

release to water; hence the subsoils will likely represent a net sink for P (McDowell and 

Sharpley, 2001). Land-disturbing activities (e.g., urban development and mining) can be 

a significant source of sediment P, particularly when eroded sediments are rich in 

nutrients because of past agricultural practices. For example, construction of one side 

channel on the Missouri River floodplain has been calculated to contribute ~ 4,000 metric 

ton P (4,400 ton P) to the river (Kristin Perry, Missouri Clean Water Commission, 

presentation to SAB Panel, June 2007). 

Regardless of sediment source, particulate P is the predominant form in transport. 

Both fluvial hydraulics and adjacent land use influence the properties of sediment within 

river systems (McDowell et al., 2002). To link P loss from the landscape to channel 

processes, variability in flow, local sources of P, sediment properties, and changes in P 

forms and loads should be simulated in models that estimate P loss from catchments, 

although this is rarely done. 

 In tile-drained agricultural regions, P is transported to streams by both overland 

flow and by the artificial drainage systems, which have been associated with elevated 

dissolved reactive P (DRP) concentrations (Xue et al., 1998). DRP concentrations 

remained high in successive tile flow events, suggesting a pool of soil P that is readily 
102

desorbed (Gentry et al., 2007). In a tile-drained Illinois watershed, P loss via tiles 

represented 45% of total P loss in one year and 91% during a wetter year (Gentry et al., 

2007). One rain-on-snow event transported about 40% of the annual P load in one week, 

80% of which was DRP (Gentry et al., 2007). Clearly artificial drainage alters both the 

amount and form of P exports, and the amount exported is dependent on both the 

magnitude and timing of storms. 

 In fluvial systems with good hydraulic mixing (e.g., shallow streams), P 

availability in sediments can be estimated by the equilibrium P concentration (EPC0). At 

low flow, EPC0 will have a major influence on soluble P concentration, for P will desorb 

from sediments if P concentration in water is less than the sediment’s EPC0, or P will 

adsorb to sediments if P concentration is greater than EPC0. P desorbed from sediments 

will be available for biological uptake. Bioavailable P from desorption is likely to be 

most significant as salinity increases in the estuary (Sutula et al., 2004). 

 Although cellular uptake and growth rates are generally saturated at low P 

concentrations, maximum biomass accrual in streams often occurs at somewhat greater 

concentrations (0.015–0.050 mg PO4-P/L, Popova et al., 2006). This range of dissolved 

P concentrations might be more typical of streams draining agricultural catchments, and 

therefore, algal and microbial uptake likely plays a significant role in dissolved P 

retention, especially at low flow. Dissolved P uptake rates of algae vary with light, water 

velocity, temperature, grazing, and time following in-stream disturbances (Mulholland et 

al., 1994). 

 Estimates of the percentage of total P inputs removed by these in-stream 

processes in regional drainages in the Mississippi River basin range from 20-75% 

(SPARROW model, Figure 38). The Ohio River basin removed the lowest percentage 

and the Arkansas River basin the highest. These percentages are considerably higher 

than what was used in the Integrated Assessment (28 to 37% in small streams and 

negligible in the mainstem). 

 P concentrations and loads generally increase with increasing discharge and are 

greatest on the rising limb of the hydrograph (e.g., Green and Haggard, 2001; Novak et 

al., 2003; Richards et al., 2001). Although P concentrations are greater during high 

flows, the importance of in–stream P retention is minimized at those times because of 

sediment resuspension and scouring within the channel. However, P deposition on 

floodplains may be a significant P sink during storms. Many streams export most of their 

P loads during episodic storm events; e.g., in Illinois agricultural watersheds, extreme 

discharges (>90th percentile) are responsible for 84% of P export and 98% of P export 

occurred at discharges > median (Royer et al., 2006). This export is primarily particulate 

P; in contrast, over half of dissolved P export can occur during base flow conditions 

(Novak et al., 2003). Dissolved P constitutes a larger proportion of P export in 

watersheds with extensive tile drainage (Royer et al., 2006). Because most P transport 

occurs at high flows, models from Illinois agricultural watersheds suggest that enhancing 

in-stream P removal by 50% during low flows (e.g., less than the median) would reduce P 
103

export by less than 1%, whereas enhancing P removal by 25% during high flows (more 

than the 75th percentile) would increase P removal by 24% (Royer et al., 2006). 

 Silicate. Understanding of Si transport and transformations in rivers and streams 

lags far behind that of N and P. Although first generation models for Si transport and 

transformations are available (Garnier et al., 2006; Sferratore et al., 2006), there are 

currently no models in the Mississippi River basin to predict the transport of dissolved 

silicate or biogenic Si (amorphous Si contained in diatoms and phytoliths). Once 

dissolved silicate is weathered, there are a number of transformations that occur including 

inorganic transformations (such as new clay formation and precipitation as amorphous Si 

in soils) and biological transformations (such as the uptake and deposition in terrestrial 

plants, uptake and deposition in diatoms in aquatic systems) (Conley, 2002). Unlike 

models developed for N and P, there are no models that describe the complexity of 

biological transformations that occur with Si. In addition, significant reductions in the 

transport of Si have occurred with the building of dams along the Mississipi River 

leading to potentially significant changes in food webs on the Mississippi River shelf 

(Turner and Rabalais, 1994). 

Freshwater wetlands 

 The Integrated Assessment recognized the historical loss of many freshwater 

wetlands as one of the primary land-use changes contributing to excess nutrient loads in 

the Mississippi River basin. Mitsch et al. (1999) suggested the creation and restoration of 

wetlands for the specific purpose of controlling non-point source nutrient loads and 

emphasized the importance of targeting wetland creation and restoration in areas where 

nitrogen concentrations and loads were highest. They estimated that restoring about 2 

million ha (5 million ac) of wetlands would reduce N loads to the Gulf of Mexico by 

20%, assuming a denitrification rate of 150 kg N/ha (134 lb N/ac) of wetland /yr. 

Subsequent research (Section 4.5.2 of this report) suggests that wetlands can achieve 

substantially higher N removal rates in areas with elevated nitrate concentrations (Figure 

12), underscoring the importance of targeting restorations. 

 Wetland restoration is a particularly promising approach for heavily tile drained 

areas like the Corn Belt (Figure 10). This region was historically rich in wetlands, and in 

many areas, farming was made possible only as a result of extensive wetland drainage 

(Dahl, 1990; Pavelis, 1987). There are widespread opportunities for wetland restoration 

in the Mississippi River basin, and since the CENR reports, approximately 570,000 ha 

(1.4 million ac) of wetlands have been restored, created, or enhanced within the basin 

under the Wetland Reserve Program (WRP), Conservation Reserve Program (CRP), 

Conservation Reserve Enhancement Program (CREP), Environmental Quality Incentive 

Program (EQIP), and Conservation Technical Assistance (CTA) (Table 7). However, the 

vast majority of wetland restorations have been motivated primarily by concern over 

habitat loss, and site selection criteria for wetland restorations have not primarily 

considered water quality functions. This past emphasis does not lessen the promise of 

wetlands for water quality improvement but rather underscores the need for programs 
104

focused on restoring wetlands explicitly for the purpose of reducing non-point source 

nutrient loads. 

Table 7: Acres of wetlands created, restored or enhanced in major subbasins of the Mississippi River from 

2000-2006 under the Wetland Reserve Program (WRP), Conservation Reserve Program (CRP), 

Conservation Reserve Enhancement Program (CREP), Environmental Quality Incentive Program (EQIP), 

and Conservation Technical Assistance (CTA). (Personal communication, Mike Sullivan, USDA). 

2-digit Watershed Hectares 

Ohio River basin 33,300 

Tennessee River basin 2,130 

upper Mississippi River basin 133,227 

lower Mississippi River basin 241,868 

Missouri River basin 93,108 

Arkansas, White, and Red River basins 68,161 

Total 571,794 

Nutrient Sources and Sinks in Coastal Wetlands 

 The general conclusion in the Integrated Assessment was that coastal wetlands are 

of secondary importance as nutrient sinks in comparison to other sources and sinks. 

Their role as a source of organic matter was discussed in an earlier section, and a more 

detailed review of that subject is in Section 2.1.5. Mitsch et al. (1999) assessed the utility 

of wetlands as nutrient and sediment sinks and concluded that 1) potential NO3 reduction 

by coastal wetlands was likely less than 10-15% of the total river load; 2) water passage 

through coastal wetlands would likely decrease water column N:P and N:Si ratios; 3) the 

concept of coastal wetlands as net nutrient sinks remains controversial (e.g., Turner, 

1999) so more large-scale measurements are needed; 4) deltaic systems might become Nsaturated or begin to release N in forms other than NO3; and 5) research and modeling 

was needed to better understand relationships between land subsidence, river diversions 

into wetlands, and N uptake in the coastal wetland/delta area. The Integrated Assessment

concluded that although coastal denitrification rates were substantial (10-25 g N/m2

/yr) 

relative to many shallow estuarine areas, diversion of river water into coastal wetlands 

might lead to N removal rates of 50-100 metric ton N/yr (55-110 ton N/yr), which is a 

relatively small fraction of N reduction goals. 

 A number of papers have been produced concerning nutrient sources and sinks in 

coastal wetlands since the Integrated Assessment. Lane et al. (2002) reported large 

decreases in nitrate as river water passed through an estuarine/wetland complex 

(Fourleague Bay); this estuarine-marsh complex appears to buffer the impact of the 

Atchafalaya River on coastal waters by causing an estimated 41to 47% reduction in river 

nitrate concentrations. Denitrification rates in coastal wetlands ranged from 30 to 40 g 

N/m2

/yr (larger than rates typically measured in adjacent estuaries), accretion rates of 8-

11 mm/yr or about 2,300 g dry sediments m2

/yr (approximating sea level rise), and N 

burial rates of about 7 g N/m2

/yr. Day et al. (2003) and others argued for river diversions 
105

to wetlands to prevent land losses and remove nutrients via denitrification, burial, and 

plant uptake. Nitrogen reductions of about 4 g N/m2

/yr and 10-20 g N/m2/yr have been 

recorded for forests and wetlands, respectively. Particulate N burial rates of 13 to 23 g 

N/m2

/yr have been measured in some wetlands. These are substantial rates by estuarine 

standards but modest relative to wetlands/reservoirs in the upper MARB connected to or 

adjacent to agricultural drainage. However, Turner (1999) reported very small N 

concentration reductions and modest TSS, POC, and particulate P concentration 

reductions in waters flowing through the Atchafalaya system and hence concluded river 

diversions would remove small amounts of nutrients relative to nutrient input loads. 

 The recent literature supports the importance of forested and other types of coastal 

wetlands for nutrient uptake and sediment accretion, both of which would lead to 

reductions in loads to the GOM. Rates appear to be substantial compared to most subtidal estuarine locations (excluding areas like the Mississippi River plume) and moderate 

to small relative to many freshwater natural and created wetlands. Rates lower than those 

observed in more northern wetlands of the MARB may be due to the generally lower 

nutrient loading rates to these coastal wetland systems (< 10 g N/m2

/yr). However, given 

the data currently available, it is doubtful that a predictive model of nutrient losses in 

these coastal wetlands can be developed following the general form of statistical models 

used for predicting nutrient losses in freshwater wetlands (Saunders and Kalff, 2001; 

Spieles and Mitsch, 2000). 

 Missing from the GOM hypoxia analysis is a regional scale (i.e., larger spatial 

scale) analysis of both nutrient (N and P) and OM losses associated with coastal 

wetlands. It would appear that sufficient information is currently available to delineate 

the spatial extent of various coastal wetland habitats. It seems less certain that essential 

nutrient and OM loss rate estimates (e.g., long-term burial of C, N and P or 

denitrification) are available to achieve this goal. 

 There appear to be few nutrient, sediment, or organic matter budgets available for 

these coastal wetlands that can be used to judge the effectiveness of wetlands as either 

sinks or sources. For example, nutrient sink behavior of wetlands has been inferred from 

nutrient concentration reductions with distance from a nutrient source. While this 

approach has appeal, it would be more convincing if nutrient loads (i.e., concentrations 

coupled to water flows) entering and leaving wetland systems were compared in a mass 

balance format. Additionally, more emphasis on process measurements (e.g., burial, 

denitrification, and plant uptake rate) would allow for better understanding of observed 

differences between wetland inputs and outputs. It appears that process measurements in 

these coastal wetlands lag behind those made in natural and created wetlands in other 

parts of the MARB. 
106

Key Findings and Recommendations 

 The percentage of annual N and P inputs removed by in-stream processes varies 

by MARB sub-basin and ranges from 20 to 55% for N and 20 to 75% for P based on 

model estimates. There currently are no models to predict the transport of dissolved 

silicate. Denitrification can be a significant pathway for N removal in small streams 

during low flow, warm periods, thereby enhancing local water quality. However, most 

nitrate is exported to the Gulf during high flows in the period from January to June, when 

denitrification is not effective in removal. Since the effectiveness of N removal in 

aquatic systems increases with water residence time, enhancing the connectivity and 

water residence time in floodplains and backwater areas on the upper Mississippi River 

during periods of high flow and high nitrate concentrations could increase the 

effectiveness of N removal by denitrification and, therefore, reduce the N flux to the 

Gulf. Likewise, since high flow events that generate overland flow in agricultural 

regions generally account for most of the annual P flux (as much as 84%, primarily as 

particulate P), deposition on floodplains and in backwater areas could represent a 

significant P sink. However, in tile drained areas, dissolved reactive P represents a much 

larger percentage of P flux (45-91%), and deposition is a less significant sink. 

 There has been substantial wetland restoration within the MARB since the CENR 

reports, but restorations have not been targeted for water quality benefits. The greatest 

water quality benefits will be realized in areas of the Corn Belt with highest nitrate 

concentrations and loads. 

 Although current estimates of denitrification rates in coastal wetlands are higher 

than the estimates used in the Integrated Assessment, current studies still conclude that 

river diversions to coastal wetlands would remove only small amounts of nutrients 

relative to the total fluxes. However, better estimates of nutrient and organic matter loss 

rates (denitrification; long-term burial of C, N, and P; and plant uptake) are needed to 

better understand observed differences between wetland inputs and outputs in coastal 

areas. Based on these findings, the SAB Panel offers these recommendations. 

• Removal of both N and P can be increased by implementing management 

strategies that include enhancing hydrologic exchange and retention on 

floodplains and in backwater habitats when discharge, total P and nitrate 

concentrations are high (e.g., during spring), particularly in rivers of 

intermediate size. 

• More reliable and process-driven models that simulate fluvial processes 

and estimate N and P transfer to stream channels need to be developed to 

more accurately predict land management or BMP impacts on nutrient 

inputs to receiving waters. 

• First-generation models need to be developed to describe the transport and 

transformations of Si in the MARB. 

• Programs focused on restoring wetlands explicitly for the purpose of 

reducing non-point source nutrient loads need to be implemented and 
107

targeted in areas of the Corn Belt with highest nitrate concentrations and 

loads. 

• Better measurements of key processes (e.g., burial, denitrification, plant 

uptake rate) are needed in coastal wetlands to provide a better 

understanding of observed differences between inputs and outputs. 

• Regional scale studies of coastal wetlands are needed to develop nutrient, 

sediment, and organic matter budgets that can be used to better evaluate 

the effectiveness of coastal wetlands as sinks or sources. 

3.4. Ability to Route and Predict Nutrient Delivery to the Gulf 

 The SAB Panel concurs with the Integrated Assessment’s identification of 

modeling as a critical component of an adaptive management approach to improving Gulf 

hypoxia. Along with monitoring, interpretation, and research, modeling can improve the 

scientific understanding of the impacts of land and nutrient management actions on 

watershed and Gulf of Mexico environmental quality. The Integrated Assessment used 

only limited modeling results at the field-scale and none at the watershed scale in its 

assessment. 

Research was proposed in the Integrated Assessment to develop an effective 

modeling framework, including improved watershed and basin-scale simulation of 

nutrient transport and transformations from natural, urban, and agricultural landscapes, 

improved estimates of nutrient mass balances throughout the landscape and improved 

understanding of biogeochemical cycling within the basin. Within this modeling 

framework, further research was called upon to assist in four areas: 1) to characterize the 

dynamics and timing of nutrient movement from the edge of the field in agricultural 

landscapes to small streams and tributaries, particularly from agricultural tile drainage 

systems; 2) to scale up from experimental plots to watershed/farm-scale studies of onfarm practices and edge-of-field strategies to reduce and intercept nutrients; 3) to assess 

the effects on nutrient loads and hypoxia of long-term change in climate, hydrology, and 

population; 4) to evaluate the role of flood events and the potential role of flood 

prevention strategies on nutrient transport to the NGOM; and 5) to improve 

understanding of the social and economic trade-offs and impacts of various management 

and policy alternative strategies. 

Numerous models have been used to describe sources, transport, and delivery of 

nutrients at various spatial and temporal scales within the MARB (Table 8). Several of 

these studies address needs identified in the Integrated Assessment, including improved 

understanding of basin-scale nutrient transport, nutrient cycling processes, tile-drainage 

nutrient transport, watershed-scale simulation of in-field and edge-of-field practices, 

climate effects, and loading from high-flow events. Issues associated with social and 

economic tradeoffs of alternative strategies are discussed in Sections 4.3 and 4.4. 
108

 While each of the models listed in Table 8 may prove useful for developing 

adaptive management to mitigate hypoxia, we single out three models for further 

discussion based on the fact that they have been applied at the basin wide scale within the 

MARB. In so doing, we do not intend to suggest that only these three models should be 

relied upon for insight into the processes of nutrient fate and delivery. 

Table 8. Attributes of models used to estimate sources, transport and/or delivery of nutrients to the Gulf of 

Mexico. 

1

NOTE: Although all models cited above (as well as numerous other models not included) are relevant to 

MARB and Gulf of Mexico hypoxia issues, not all of the above models have been discussed within 

this report. In this section, the discussion was focused on models considered most applicable to 

MARB basin-scale processes. More details on applications of these models within the MARB can be 

found in cited references. 

2

Type—model classification: S=statistical/stochastic, C=export coefficient, B=mass balance, 

E=empirical/process-based, M=mechanistic/process-based 

3

Time—smallest time-scale for output: S=subdaily, D=daily, M=monthly, A=annually 

4

Space—organizing spatial scale: F=field, H=hydrologic resource unit, W=watershed, B=basin, R=river 

network 

5

Components—model features: R=runoff, D=drainage, S=snowmelt, E=erosion, N=nutrients, Q=stream 

processes, P=ponds/reservoirs, U=urban 

6

Inputs—input types: C=climate (temperature, precipitation), F=fertilizers, P=point sources, 

A=Atmospheric deposition, S= spatial land-use, L=single-field land use, G=stream gage data 

7

Outputs—constituents modeled: F=flow, S=sediment, N=nitrogen, P=phosphorous, C=carbon 

Model1 Type2 Time3 Space4 Components5 Inputs6 Outputs7 Predicts8 Strength9 MARB 

Refs10

ADAPT E,M D F,W R,D,S,E,N,P,U C,F,P,A,S F,S,N,P C,L,M Overland/drainage 1 

AnnAGNPS E,M D H,W R,E,N,P,U C,F,P,A,S F,S,N,P C,L,M In-field sediment 2 

DAFLOW/BLTM M D R R,N,Q C,G F,N C,L River routing 3 

DRAINMOD E,M D F R,D,N C,F,L F,N C,L,M Drainage 4 

EPIC E,M D F R,S,E,N C,F,A,L F,S,N,P C,L,M In-field practices 5 

GLEAMS E,M D F R,E,N C,F,A,L F,S,N,P C,L,M Overland 6 

HSPF, LSPC E D W R,D,S,E,N,Q,U C,F,P,A,S F,S,N,P C,L,M,R Overland/stream -- 

IBIS/THMB E,M D W,B R,D,S,E,N,Q,P C,F,P,A,S F,S,N,C C,L,M Ecosystem 8 

L-THIA C,E A F,W R,E,N,U C,P,N,F F,S,N,P L W-s land-use 

change 

-- 

NANI B A W,B N,P F,P,A,S N C,L,M Process accounting 10 

PLOAD C M W R,E,N,U F,C,P,S F,S,N,P C,L,M Distribute w-s loads -- 

REMM M D F R,D,E,N C,F,L F,S,N,P,C C,L,R Riparian ecosystem 12 

RZWQM E,M S F R,D,S,N C,F,A,L F,S,N,P C,L,M Subsurface/plants -- 

SPARROW S,M A W,B R,D,N,Q,P,U C,F,P,A,S,G F,N,P C,L Data-driven 14 

SWAT E,M S,D H,W,B R,D,S,E,N,Q,P,U C,F,P,A,S F,S,N,P,C C,L,M,W,R Overland/w-s 15 

WARMF E,M D W R,E,N,P,U C,F,P,A,N F,S,N,P C,L,M TMDL study -- 

WEPP M S F,W R,D,S,E,N C,F,P,A,S F,S,N,P C,L,M,R Hillslope 17 
109

8

Predicts—predictive capability: C=climate change, L=land-use change, M=land-management change, 

W=wetland change, R=riparian change 

9

Strength—application for which model tends to be well suited (w-s=watershed) 

10MARB Refs—Recent and key references from Mississippi/Atchafalaya River basin model applications: 

-------------------------------------------------------------------------------------------------------------------------------------- 

1—ADAPT: Dalzell, B.J., P.H. Gowda, D.J. Mulla. 2004 

Gowda, P.H., D.J. Mulla. 2006. 

Gowda, P.H., B.J. Dalzell, D.J. Mulla. 2007 

Sogbedji, J.M., G.F. McIsaac. 2006. 

2—AGNPS: 

Yuan, Y., R.L. Bingner, R.A. Rebich. 2001 

Yuan, Y., R.L. Bingner, F.D. Theurer. 2006 

Yuan, Y. R.L. Bingner, F.D. Theurer, R.A. Rebich, P.A. Moore. 2005 

3—DAFLOW/BLTM: 

Broshears, R.E., G.M. Clark, H.E. Jobson. 2001 

4—DRAINMOD: 

Northcott, W.J., R.A. Cooke, S.E. Walker, J.K. Mitchell, M.C. Hirschi. 2001 

5—EPIC:

Atwood, J.D., V.W. Benson, R. Srinivasan, C. Walker, E. Schmid. 2001 

Chung, S.W., P.W. Gassman, R. Gu, R.S. Kanwar. 2002

6—GLEAMS: 

Wedwick, S., B. Lakhani, J. Stone, P. Waller, J. Artiola. 2001 

8—IBIS/THMB:

Donner, S.D., M.T. Coe, J.D. Lenters, T.E. Twine, J.A. Foley. 2002 

Donner, S.D., C.J. Kucharik. 2003a 

Donner, S.D., C.J. Kucharik, J.A. Foley. 2004. 

Donner, S.D. 2006 

10—NANI:

McIsaac, G.F., M.B. David, G.Z. Gertner, D.A. Goolsby. 2002 

Howarth, R. W., G. Billen, D. Swaney, A. Townsend, N. Jarworski, K. Lajtha, J. A. Downing, R. Elmgren, N. 

Caraco, T. Jordan, F. Berendse, J. Freney, V. Kueyarov, P. Murdoch, and Zhu Zhao-liang. 1996. 

12—REMM:

Graff, C.D., A.M. Sadeghi, R.R. Lowrance, R.G. Williams. 2005 

14—SPARROW:

Alexander et al., in press. 

Alexander, R.B., R.A. Smith, G.E. Schwarz., 2004. 

Smith, R.A., G.E. Schwarz, R.B. Alexander, 1997 

Alexander, R.B., R.A. Smith, G.E. Schwarz., 2000 

15—SWAT: 

Anand, S., K.R. Mankin, K.A. McVay, K.A. Janssen, P.L. Barnes, G.M. Pierzynski. 2007 

Du, B., A. Saleh, D.B. Jaynes, J.G. Arnold. 2006 

Gassman, P.W., M.R. Reyes, C.H. Green, J.G. Arnold. 2007 

Green, C.H., M.D. Tomer, M. DiLuzio, J.G. Arnold. 2006a 

Hu, X., G.F. McIsaac, M.B. David, C.A. Louwers. 2007 

Jha, M., J.G. Arnold, P.W. Gassman, F. Giorgi, R.R. Gu. 2006 

Kirsch, K., A. Kirsch, J.G. Arnold. 2002. 

Santhi, C., J.G. Arnold, J.R. Williams, W.A. Dugas, R. Srinivasan, L.M. Hauck. 2001 

Shirmohammadi A., I. Chaubey, R. D. Harmel, D.D. Bosch, R. Muñoz-Carpena, C. Dharmasri, A. Sexton, M. Arabi, 

M.L. Wolfe, J. Frankenberger, C. Graff, T.M. Sohrabi. 2006.. 

Stone, M.C., R.C. Hotchkiss, C.M. Hubbard, T.A. Fontaine, L.O. Mearnes, J.G. Arnold. 2001 

Vache, K.B., J.M. Eilers, M.V. Santelman. 2002 

VanLiew, M.W., T.L. Veith, D.D. Bosch, J.G. Arnold. 2006 

Wang, X., A.M. Melesse. 2005 

17—WEPP:

Tiwari, A.K., L.M. Risse, M.A. Nearing. 2000 
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SPARROW Model 

 The SPARROW (SPAtially Referenced Regressions On Watershed attributes) 

model is a hybrid mechanistic/empirical, basin-scale simulation model developed by U.S. 

Geological Survey (Smith et al., 1997; Alexander et al., 2007a). The model uses spatially 

distributed data on nutrient sources, climate, soils, topography, and natural and artificial 

drainage densities to estimate N and P delivery to streams and removal processes in 

streams and reservoirs under long-term steady-state conditions. Nutrient sources include 

atmospheric deposition (N only), urban/human sources, agricultural runoff and 

subsurface drainage, and natural sources from forest, barren, and shrub lands. 

 Wet-deposition data are the basis for atmospheric deposition N source, which 

assumes that dry deposition and ammonium deposition are spatially correlated to wet 

deposition. Urban nutrient sources include all human-population-dependent nutrient 

sources: municipal and septic-system wastewater, stormwater runoff, and other sources 

that are spatially correlated to human population data (such as wet and dry deposition 

from vehicles, power plants, etc.). Agricultural nutrient sources include commercial 

fertilizers, livestock manure, and biological N2 fixation. Soil transformations of N and P 

are not considered and assumed to be in equilibrium between immobilization and 

mineralization. 

SPARROW simulates N and P fluxes (mass) and yields (mass per unit area) 

within sub-catchments using three first-order attenuation terms, with mass-balance 

constraints, to represent nutrient losses in overland transport, riverine processes, and 

reservoir trapping. Model parameters are calibrated based on the source conditions for a 

base year and the flow-adjusted, long-term mean annual loads of total N and total P 

estimated using rating curves fit to stream monitoring data. The calibrated SPARROW 

model can be used to assign these loads to specified nutrient sources and sub-catchments 

with quantifiable uncertainty. 

SPARROW has been applied to diverse watersheds including the MARB (Smith 

et al., 1997; Alexander et al., 2000; Alexander et al., 2007a), the Chesapeake (Preston 

and Brakebill, 1999), the Neuse (McMahon et al., 2003) in North Carolina, and the 

Waikato (Alexander et al., 2002a) in New Zealand. The most recent application of 

SPARROW to the MARB includes more and better data for model parameter estimation 

and greater detail in the model specification (Alexander et al., in press). The result is an 

increased number of nutrient-source terms and 20% less model error compared to 

previous applications of SPARROW (Smith et al., 1997; Alexander et al., 2000). Several 

important assumptions are embedded in the modeling approach, however, and these must 

be considered in interpretation of model results for the MARB. 

SPARROW does not assess flow-related changes in nutrient loads, which are 

important to Gulf hypoxia extent and severity. In comparing SPARROW predictions 

with observed loads for a particular period or location, SPARROW does not estimate 

nutrient load for any particular year, but rather a flow-adjusted or flow-independent load. 
111

This is the load predicted under long term average flow conditions for the source input 

conditions of a particular year (the base year). For example, the SPARROW estimates 

for the 1992 base year in Alexander et al. (in press) are the mean annual loads that would 

be predicted under the source conditions of 1992 and the mean annual flow of the period 

from 1975-2000. These represent the loads that would have been expected in 1992 if 

1992 had had the mean annual flow of the period 1975-2000. They are not an estimate of 

the nutrient loads in1992. As a result, the comparisons between 1992 and 2002 in 

Alexander et al. (in press) are not based on the loads predicted for 1992 and 2002 but 

rather the loads that would have been expected in 1992 and 2002 under the different 

source inputs for those two years, but assuming both years had had exactly the same flow 

patterns, i.e. the mean annual flow of the period 1975-2000. 

Model input coefficients for each nutrient source are statistically estimated by the 

model, and as such, are influenced by all sources that are spatially correlated to these 

sources (whether these correlated sources are in the model or not). For example, wetdeposition N was spatially characterized from monitoring data, but no data for dry 

deposition were used; and urban sources were modeled assuming that all sources were 

correlated and spatially distributed similar to the model input of population. Particularly 

in these two cases, coefficients may be artificially high as they include the effects of other 

spatially correlated sources that are not in the fitted model. The model does not account 

for soil storage of nutrients, but assumes that stream inputs are correlated to the 

agricultural nutrient source inputs. The lack of a soil storage term may ignore nutrient 

carry-over effects that are often important in determining stream export (David et al., 

1997; David and Gentry, 2000; McIsaac et al., 2001; Mulvaney et al., 2001). These 

limitations and others discussed in Alexander et al. (in press) should be considered in 

interpretation of the SPARROW model results. 

The most recent version of SPARROW (Alexander et al., in press) is thought to 

have improved many aspects of this statistical model. For example, the percentage of N 

and P that enters streams and is actually delivered to the NGOM has increased, as instream removal terms have been reduced. SPARROW can be used to examine source 

inputs for the nutrients being transported by streams, and with each new version these can 

change. Source areas are quite dependent on land-to-water transfer coefficients, and the 

way the model represents inputs and their availability to be transferred to a stream. As 

expected, agriculture was found to be the major source of nutrients to the NGOM in this 

recent application (Alexander et al., in press). Important non-agricultural nutrient 

contributions were from atmospheric deposition and urban sources. The largest source of 

N is attributed to fertilizer inputs to corn and soybean fields (52%) followed by 

atmospheric deposition (16%). In contrast, the largest source of P is attributed to animal 

manure on pasture and rangelands (37%) followed by corn and soybeans (25%), other 

crops (18%), and urban sources (12%) (Alexander et al., in press). It is important to note 

that in the model structure, manure is the only source of P that is available for transport 

from pasture and rangelands. These lands are otherwise assumed to be in steady state. 

Similarly, fertilizer, N fixation, and manure N are the only source of N that is available 

for transport from corn and soybean. These lands are assumed to be in long term steady 

state, and there is assumed to be no net soil mineralization. Statistical coefficients for 
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agricultural sources suggested N delivery to streams ranging from 6% of applied nutrients 

for pasture/rangeland to 16% for corn and soybeans, and the opposite trend for P 

delivery, ranging from 2% for corn and soybeans to 14% for pasture/rangeland. These 

results give a very different picture of important inputs to the basin and their effects on 

riverine N and P fluxes. For example, atmospheric deposition (for N) and manure (for P) 

are thought to much more important, and point sources and corn and soybean production 

much less important than mass balance type calculations would suggest (see Section 3.2). 

SPARROW is continually being developed and improved. However, the Panel cautions 

against the sole use of SPARROW (or any model) for making decisions about where to 

target management efforts given the current stage of development of this approach. 

SWAT Model 

 The Soil and Water Assessment Tool (SWAT) model is a physically based, 

deterministic, continuous, watershed-scale simulation model developed by the USDA 

Agricultural Research Service (Neitsch et al., 2004; Arnold et al., 1998). It uses spatially 

distributed data on topography, soils, land cover, land management, and weather to 

predict water, sediment, nutrient, and pesticide yields. A modeled watershed is divided 

spatially into subwatersheds using digital elevation data according to the density 

specified by the user. Subwatersheds are modeled as having uniform slope and climatic 

conditions, and they are further subdivided into lumped, nonspatial hydrologic response 

units (HRUs) consisting of all areas within the subwatershed having similar soil, land 

use, and land management characteristics. The use of HRUs allows soil and land-use 

heterogeneity to be simulated within each subwatershed but ignores pollutant attenuation 

between the source area and stream and limits spatial representation of wetlands, buffers, 

and other BMPs within a subwatershed. 

The model includes subbasin, reservoir, and channel routing components. The 

subbasin component simulates runoff and erosion processes, soil water movement, 

evapotranspiration, crop growth and yield, soil nutrient and carbon cycling, and pesticide 

and bacteria degradation and transport. It allows simulation of a wide array of 

agricultural structures and practices, including tillage, fertilizer and manure application, 

subsurface drainage, irrigation, ponds and wetlands, and edge-of-field buffers. The 

reservoir component detains water, sediments, and pollutants, and degrades nutrients, 

pesticides and bacteria during detention. The channel component routes flows, settles 

and entrains sediment, and degrades nutrients, pesticides and bacteria during transport. 

SWAT typically produces daily results for every subwatershed outlet, each of which can 

be summed to provide monthly and annual load estimates. 

The SWAT model has been tested for a wide range of regions, conditions, 

practices, and time scales (Gassman et al., 2007). Evaluation of monthly and annual 

streamflow and pollutant outputs indicate SWAT functioned well in a wide range of 

watersheds. Relatively poor results in some cases, particularly for daily flow and 

pollutant outputs, were attributed partly to input and calibration data uncertainty and 

partly to model limitations. In general, the model had more difficulty simulating wet 
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years than dry years and tended to overestimate soil water in dry soil conditions and 

underestimate in wet soil conditions. 

Numerous studies have applied the SWAT model in the Mississippi River basin. 

Several recent studies have addressed issues identified in the Integrated Assessment, such 

as application of field-scale hydrologic processes to large watershed scale (Arnold et al., 

1999; Anand et al., 2007), effectiveness of various nutrient-reduction strategies in 

agricultural watersheds (Santhi et al., 2001; Vache et al., 2002; Hu et al., 2007), model 

enhancements to address tile-drained cropland (Du et al., 2006; Green et al., 2006a), and 

assessment of the impacts of climate change on large-basin hydrology and nutrient export 

(Jha et al., 2006). These studies are discussed below. 

Studies from field scale (Anand et al., 2007) to basin-scale (Arnold et al., 1999) in 

the MARB have demonstrated the ability of SWAT to scale-up processes to the largewatershed scale. Arnold et al. (1999) validated the water-balance component of SWAT 

in a large-scale modeling study of the conterminous U.S. and concluded that it would be 

useful in studying the effects of climate and BMPs on annual and seasonal runoff. The 

long-term effects of various BMPs was assessed in a 4,277 km2

 (1,651 mi2) pasturerangeland-dominated watershed experiencing urban growth in Texas (Santhi et al., 2001). 

They found future (2020) loads could be reduced by about 50% by implementing a 

combination of practices, including a 1 mg/L limit for wastewater treatment plant P 

effluent, limiting dairy manure land applications to the P rate, exporting 38% of manure 

from the watershed, and reducing P in livestock diet. Vache et al. (2002) explored the 

impact of traditional and alternative agricultural practices on water quality in two 

agricultural watersheds in Iowa. Continuing current trends in Midwestern agricultural 

production to 2025 (including increased conservation tillage, increased farm size and 

total acres, and current BMPs) resulted in simulated increase of nitrate export and 

decrease of sediment export relative to present. Two other scenarios representing 

different combinations of practices, such as complete conversion of cropland to no-till, 

implementation of riparian buffers on all streams, and increased use of perennial cover 

(CRP, pasture, and alfalfa), resulted in reductions of nitrate loads by 54 to 75% and 

sediment load by 37 to 67% simulated by SWAT. In a tile-drained watershed in eastcentral Illinois, reductions in N fertilizer resulted in 10 to 43% decrease in riverine nitrate 

export (Hu et al., 2007). However, SWAT overestimated nitrate export during major wet 

periods and had several other unrealistic aspects of N cycle components. Recent 

enhancements have been made to allow better simulation of tile-drainage in agricultural 

fields by SWAT (Du et al., 2006; Green et al., 2006a). This change indicates that 

previous modeling results by SWAT in heavily tile-drained watersheds should be 

reassessed using the revised model. 

Shifts in future precipitation and climate may impact flow and nutrient loads from 

the MARB. Jha et al. (2006) used SWAT to assess the effects of future climate change on 

UMRB flows. They found a doubling in CO2 (to 660 ppmv) to result in a 36% increase 

in average annual streamflow and a 20% increase in precipitation to increase streamflow 

by 58%. Similar increases were found in average monthly streamflow in the April to 

May period, which is in the critical period for hypoxia development. Mean annual 
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streamflow changes in response to six general circulation model scenarios ranged from 

-6% to +51%. Results indicated increases in rainfall and snowmelt in January and 

February and large increases in spring stream flow. 

IBIS/THMB Model 

The Integrated Biosphere Simulator (IBIS) land-surface and terrestrial ecosystem 

model and the Terrestrial Hydrology Model with Biogeochemistry (THMB, an enhanced 

version of Hydrological Routing Algorithm, HYDRA) are two physically based models 

that have been linked to model large basin-scale hydrological, carbon, and nutrient 

processes (only nitrogen at this time) (Donner et al., 2002; Donner, 2006). The IBIS 

model represents phenomena such as land-surface biophysical processes, canopy 

physiology, vegetation phenology, and long-term ecosystem dynamics at different time 

steps, ranging from 60 minutes to 1 year, to simulate time-transient surface and 

subsurface hydrological fate and transport processes. IBIS requires spatially distributed 

inputs of climate, soil texture, vegetation type and associated management information. 

It uses these inputs to simulate terrestrial processes at a user-defined grid-cell scale. The 

terrestrial model is coupled with THMB, which represents phenomena such as solute 

transport, surface and subsurface leaching, point-source inputs, and in-stream chemical 

and biological transformations to simulate river, wetland, lake, and reservoir flow and 

storage of water and nutrients. 

The IBIS/HYDRA model, with a simplified nitrogen leaching algorithm, was 

found to represent much of the spatial and temporal variability in stream discharge and 

nitrate export within the MARB (Donner et al., 2002). A study of 29 stations in the 

MARB from 1965-1994 found interannual errors in simulated river discharge were less 

than 20% for the majority of the data (76%), although the seasonal errors were greater 

than 20% for 65% of the station months and particularly underestimated the magnitude of 

spring discharge. A similar analysis found simulated annual mean nitrate export of the 

Mississippi River at St. Francisville was within 1% of the USGS estimate, but annual 

errors at various stations varied widely. 

Results of the IBIS/HYDRA modeling study indicated that nitrate export from the 

MARB was significantly greater during the latter half of the 1955-to-1994 period, largely 

due to the increase in N fertilizer application, with greatest contribution from the central 

and eastern subbasins (Donner et al., 2002). This analysis made many simplifying 

assumptions about nitrogen inputs, fate and transport in order to isolate the impact of 

hydrology on nitrate export variability. Donner et al. (2002) concluded that the observed 

increase in river discharge was responsible for about 25% of the increase in nitrate export 

between 1966 and 1994, with an error of 7%. The remainder of the increase was inferred 

to arise predominately from an increase in fertilizer N inputs. In the Upper MARB 

(1974-1994), Donner and Kucharik (2003) found that a +/–30% change in N fertilizer 

application resulted in little change in corn yields (+4%/–10%) but greater sensitivity in 

dissolved inorganic N subsurface drainage (+53%/–37%). They note that soil N storage 

resulting from the +30% fertilizer N case appeared to lead to almost 60% increase in 

nitrate export after 20 years and that this effect was greatest during wet years. 
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Further work for the entire MARB based on IBIS/HYDRA results led Donner et 

al. (2004) to conclude that the doubling of nitrate export to the Gulf of Mexico over the 

1960-to-1994 period resulted largely from an increase in fertilizer application rates, 

particularly to corn, an increase in runoff across the basin, and the expansion of soybean 

cultivation. Their results indicated that by the 1990s, fertilized cropland (particularly in 

Corn Belt hot spots across Iowa, Illinois, and Indiana) became the overwhelming nitrate 

source in the river system, contributing almost 90% of the nitrate from just 20% of the 

watershed area. Changes in MARB crop production systems associated with a shift away 

from meat production were simulated by IBIS/THMB (Donner, 2006). Results indicated 

a reduction in total land and fertilizer demands by over 50% and N export by 49-54% 

without any change in total production of human food protein. 

Discussion and Comparison of Models 

The SAB Panel found only one study that compared any of the three focus 

models. A study comparing SWAT and a statistical approach based on SPARROW 

within the Great Ouse watershed in the United Kingdom found similar total oxidized 

nitrogen load estimations and similar statistical reliability of the two models (Grizzetti et 

al., 2005). They suggested using SPARROW as a screening tool for identifying sources 

and using SWAT for testing management practice scenarios but found that both models 

demonstrated utility for nitrogen load estimation. 

Different modeling approaches resulted in different assessments of nutrient 

sources and distribution within the MARB among the models, where comparisons were 

possible. Cropland was found to contribute 90% of nitrate within the MARB by 

IBIS/HYDRA (Donner et al., 2004) compared to 66% of total N for all crops by 

SPARROW (Alexander et al., in press). Both models suggested the major nutrient source 

yields (mass per unit area) originated from the central Mississippi and Ohio River basins. 

Future work with SWAT (J. Arnold, personal communication) applied to the entire 

MARB will provide another estimate of nutrient sources and distribution to assist with 

watershed planning and management decisions. 

Targeting 

The models cited in Table 8 vary considerably in type, scale, and approach. The 

Gulf hypoxia issue requires a diversity of model types, scales, and approaches. Models 

that can support adaptive management of Gulf hypoxia and within-region water quality 

are those that can best inform targeting of the most effective actions at the lowest cost. 

Three forms of targeting are especially important: 

• targeting sub regions or watersheds (of perhaps the 8 or 12 digit HUC size) 

that have a disproportionate effect on hypoxia and local water quality; 

• targeting the type and placement of conservation practices within those 

watersheds to achieve the greatest gains at the lowest cost; and 
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• targeting the timing of nutrient flows to best attenuate the hypoxic zone. 

Both SWAT and IBIS/THMB models directly address targeting of practices by 

simulating the effects of farm/plot scale best management practices (BMPs) directly, 

whereas the absence of BMP simulation capability is a weakness in SPARROW 

(acknowledged in Alexander et al., in press). Simulation of these BMPs is vital to the 

evaluation of successful management of nitrogen and phosphorus runoff in MARB. 

SWAT and IBIS/THMB include these practices but would benefit from additional 

verification that their mechanistic characterizations represent the range of field and 

watershed-scale processes present in the MARB. SPARROW shortcomings in 

simulation of BMPs are being addressed by ongoing efforts to evaluate farm and plot 

scale BMPs by USGS and others using data in the same sense as SPARROW is fitted 

(e.g., identifiability). When coupled with SPARROW, the results should yield a useful 

predictive model for the impact of BMP and other practices within MARB on Gulf 

hypoxia and should include uncertainty analysis. 

 Practices should be evaluated both for their impacts on total annual or long-termaverage nutrient loads as well as loads on a seasonal or other short-term time frame. 

Within-year timing of pollutant loads is simulated by SWAT and IBIS/THMB (both 

models operate on a daily basis; see Table 8) but not the annual-based SPARROW. 

Timing issues are critical for addressing seasonal water quality concerns both locally as 

well as for the April-to-June loads that appear to govern hypoxia development. 

All three models address spatial targeting of sources and implementation. 

SPARROW spatial resolution is tied to the resolution of available monitoring data; recent 

studies of the MARB have been conducted at the HUC8 level. SWAT has been applied 

at a range of watershed scales from collections of fields to Mississippi River basin, 

whereas IBIS/THMB tends to be most applicable at the larger watershed scales. Both 

SWAT and IMIB/THMB spatial resolutions currently are dictated by computing capacity 

for larger-scale basins. Spatial-scale issues are critical in targeting implementation 

actions, funding, and resources to areas with the greatest potential for improvement. 

Model interpretation must consider the relationship of watershed spatial heterogeneity 

and model averaging of input variables, process algorithms, and outputs. All three 

models provide information to assist with spatial targeting of actions. 

Finally, there is a need to integrate watershed models with economic models in 

making targeting recommendations. Due to the ability to assess the effectiveness of 

specific conservation practices in a sub watershed context, integrated economicwatershed models have largely relied on mechanistic models such as EPIC and SWAT. 

More discussion of these integrated models can be found in Section 4.3. 

Model Uncertainty 

Model predictions all have a degree of uncertainty, which should be addressed in 

presenting model results. Model uncertainty may be due to variability, inaccuracy, or 
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inappropriateness of multiple factors: 1) model algorithms or methods, 2) inputs (known 

or measured values, such as climate data), 3) parameters (values estimated based on 

functional relationship with known inputs, such as soil hydraulic conductivity or runoff 

curve number), 4) calibration data (including measurement errors associated with 

streamflow estimation and sample collection, storage, and analysis), 5) boundary 

conditions (such as initial soil moisture), 6) temporal scale (such as rainfall intensity), 

and 7) spatial scale (such as topography) (Shirmohammadi et al., 2006). Model 

uncertainty can be assessed by describing the impact of uncertain inputs and parameters, 

treated as random variables, on output variability. 

Uncertainty varies among models and differs among watersheds, depending on 

availability of data, appropriateness of model assumptions for the given watershed and 

climate, and skill of the modeler in applying the model and interpreting the results. One 

study evaluated the impact of input uncertainty on SWAT2000 model output variation 

and found that input uncertainty was transferred nonlinearly through the model 

(Shirmohammadi et al., 2006). Coefficients of variations (CVs) of 34 input parameters 

ranging from 10 to 76% resulted in a single-year output CV of only 28% for streamflow, 

lower CVs for ammonium and organic N (6-7%) and mineral P (12%), and higher CVs 

for nitrate (101%), organic P (58%), and sediment (36%). Measured streamflow, nitrate, 

and ammonium values were within one standard deviation (SD) of the mean modeled 

output, whereas sediment was within 1.5 SD. 

An advantage of SPARROW (and other models that are parameterized using 

optimization criteria, such as least squares or maximum likelihood) is that the model 

provides error terms for prediction with respect to parameter uncertainty. Process 

(mechanistic) models, such as SWAT and IBIS/THMB, are “overparameterized,” which 

means that the observational data are insufficient to provide unique/optimal estimates of 

model parameters. Thus, these process models typically have been parameterized 

according to modeler’s best judgment, with important ramifications on model 

uncertainty. New approaches acknowledge that many different parameter sets may fit 

equally well for these mechanistic and process-based models (e.g., Beven, 2001). This 

approach avoids the difficult question of how the modeler chooses an optimal, single set 

of parameters. 

Key Findings and Recommendations 

Interactions of climate, land, water body, and management factors on nutrient 

yields and loads are incredibly complex. As such, management decisions should always 

consider multiple models with different modeling approaches. The models discussed in 

this report are capable of nitrogen and phosphorus load estimation on the scale of the 

MARB, yet each has strengths and weaknesses. Other models and modeling approaches 

also exist, and each has inherent strengths, limitations, and value to improving 

understanding of and informing decision-making related to the MARB and Gulf hypoxia. 

Thus, a diversity of models needs to be developed and applied for load estimation, BMP 

evaluation, implementation targeting, and forecasting. Models should provide 
118

information about the direction, magnitude, and uncertainty of the impact of current and 

planned actions on ecosystem services at the appropriate temporal and spatial scale and at 

a resolution and precision that is appropriate to guide these decisions. With an enhanced 

modeling toolbox at their disposal, decision makers will need to select the model or 

models best suited to answering their questions and guiding their decisions. 

The uncertainty of results for each model reflects the uncertainty of the model 

structure and algorithms, as well as that propagated by the input data, user 

parameterization, calibration process, and other user-defined conditions. Other than the 

model itself, each of these factors is influenced by the skill of the model user, making it 

difficult to make blanket generalizations about reliability or applicability of the models 

discussed. 

Adaptive management will be more informative, particularly in the initial years 

post implementation, if monitoring data are used to improve models for the next iterative 

prediction. This requires that the monitoring be designed, at least in part, for this task. 

These monitoring data will also enhance the modeling effort. Rigor of model validation 

can be assessed through statistical comparison of calibration data with validation data 

provided through monitoring. Greater availability of monitoring data will allow a greater 

difference between calibration and validation data sets and provide a more rigorous 

model validation. For example, applying a model to a different watershed with different 

climate will better test the robustness of the model than a validation using a different 

period of climate data within the same watershed. 

Adaptive management will require modeling flexibility as well as consideration of 

the compatibility between watershed models, economic models, and Gulf of Mexico 

hypoxia models. The various models need to have the capability to translate across 

temporal and spatial scales and to communicate how factors affecting ecosystem services 

are simulated in order to have a smooth interface. For example, watershed model output 

of total N at annual scales should be able to interface with a Gulf model requiring daily 

conditions of inorganic N. In addition, models need to be developed and used to assess 

effects of policy decisions and management practices. Characterization of the degree of 

uncertainty would assist interpretation of results and application of these results within an 

adaptive management framework. Based on these findings, the SAB Panel offers the 

following recommendations. 

• A diversity of watershed modeling approaches, ranging from simple forecasting 

to complex statistical and mechanistic approaches, will be useful for describing 

loading and timing of nutrients to the NGOM. 

• Model selection should depend on the question(s) being asked and the associated 

strengths and weaknesses of the various models; for Gulf hypoxia, watershed 

models should address issues of management option selection, spatial targeting of 

actions, and temporal delivery of nutrient loads to the NGOM. 

• Water-quality monitoring and the documentation of critical ancillary information 
119

(i.e., inputs and management practices) should be designed, at least in part, to 

support model use and assessment, and adaptive management. 

• Uncertainty of model results should be assessed and reported. As much as 

possible, all potential sources of error should be explicitly recognized and 

discussed in reporting results. Further, confidence bounds should be reported for 

all applicable sources (e.g., parameter uncertainty). For those sources for which 

formal confidence intervals cannot be computed, sensitivity analysis or another 

form of uncertainty analysis should be undertaken and reported. 
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4. Scientific Basis for Goals and Management Options 

4.1. Adaptive Management 

Adaptive management offers a way to address the pressing need to take steps to 

manage for factors affecting hypoxia in the NGOM in the face of uncertainties. The 

authors of a recent study undertaken by the National Research Council of the National 

Academy of Sciences identified six elements of adaptive management that are directly 

relevant to goal setting and research needs (National Research Council, 2004): 1) 

resources of concern are clearly defined; 2) conceptual models are developed during 

planning and assessment; 3) management questions are formulated as testable hypotheses 

to guide inquiry; 4) management actions are treated like experiments that test hypotheses 

to answer questions and provide future management guidance; 5) ongoing monitoring 

and evaluation is necessary to improve accuracy and completeness of knowledge; and 6) 

management actions are revised with new cycles of learning. 

 Perhaps the most important “take-home” lesson from their work is contained in 

the following statement: 

Adaptive management does not postpone actions until “enough” is known about a 

managed ecosystem (Lee, 1999), but rather is designed to support action in the 

face of the limitations of scientific knowledge and the complexities and stochastic 

behavior of large ecosystems (Holling, 1978). Adaptive management aims to 

enhance scientific knowledge and thereby reduce uncertainties. Such 

uncertainties may stem from natural variability and stochastic behavior of 

ecosystems and the interpretation of incomplete data (Parma et al., 1998; Regan et 

al., 2002), as well as social and economic changes and events (e.g., demographic 

shifts, changes in prices and consumer demands) that affect natural resources 

systems. 

Thus adaptive management provides an appropriate way for decision makers to deal with 

the uncertainties inherent in the environmental repercussions of prescribed actions and 

their influences on hypoxia. 

 Adaptive management can be conducted at the several management scales that 

occur in the NGOM and MARB. On the basin scale, adaptive management requires 

measurements of both nutrient loadings and hypoxia extent (area). Although it will not 

be possible to relate these changes to specific changes in the basin, these data will 

provide better understanding of the relationships between nutrients and hypoxia. On 

smaller scales, specific management actions can be treated as experiments that test 

hypotheses, answer questions, and thus provide future management guidance at that scale 

(for example, small watersheds). 

 

 The adaptive management approach requires that conceptual models are 

developed and used and relevant data is collected and analyzed to improve understanding 

of the implications of alternative practices (e.g., Ogden et al., 2005). To help illustrate 
121

what is meant by a conceptual model, the SAB Panel has developed a diagram that shows 

major factors that affect hypoxia in the NGOM (Figure 40). The corresponding 

Figure 40: A Conceptual Framework for Hypoxia in the Northern Gulf of Mexico.
122

conceptual model would estimate the relative contribution of each influence. Those 

estimates could serve as hypotheses of relative effects, and the diagram could illustrate 

hypothesized interactions and feedbacks. Such a conceptual model organizes how 

adaptive management research is conducted in a framework where the testing of 

hypotheses and the new knowledge gained is then used to drive management adaptations, 

new hypotheses and the collection of new data on endpoints. Unlike the traditional 

model of hypothesis driven research, adaptive management implies coordination with 

stakeholders and consideration of the economic and technological limitations on 

management. Unlike traditional demonstration projects, adaptive management implies an 

understanding that complex problems will require iterative solutions that will only be 

possible through generation of new knowledge as successive approximations to problem 

solving are attempted. 

 Successful implementation of the adaptive management process is occurring in 

the Grand Canyon (Meretsky et al., 2000) and the Everglades (Sklar et al., 2005). In 

addition, steps toward adaptive management are being examined in the Upper Mississippi 

River basin (O’Donnell and Galant, in press). That work documents the need for greater 

collaboration between scientists and management agencies to plan, design, and monitor 

river enhancement programs. Problems exist in setting quantifiable success criteria, 

developing appropriate monitoring designs, and disseminating information. The SAB 

Panel expects similar difficulties in implementing adaptive management to occur 

elsewhere in the MARB. 

 There needs to be a better understanding of the spatial and temporal aspects of 

basin-level responses to management practices and also a focus on other scales at which 

response can occur in a more timely fashion. Yet observations of a basin-level response 

to practices cannot be expected for some time, which calls for management and 

evaluation to be focused on a sub-basin scale. Therefore it is important to obtain 

information at a scale where practices can be broadly and appropriately applied and 

where results are “meaningful and interpretable.” The relevant scale would likely be at 

smaller sub-watershed scales, where local water quality and quantity benefits may 

become evident more quickly. Furthermore, the demonstration of adaptive management 

within a small sub-watershed may enhance practice adoption at other locations. Thus 

conceptual models need to be developed for this scale of resolution as well. Focus at the 

small-watershed scale will also provide local water quality and quantity benefits. The 

results from small watershed studies must be able to be extrapolated to other small 

watersheds in the sub-basin and, preferably, the entire MARB, if they are to be useful in 

reducing hypoxia in the NGOM. 

 Experiments that could be applied at small watersheds to help improve 

understanding of the effects of different practices have the following characteristics: 

• Practices applied on the small watersheds should conform to accepted 

practice standards or make specific modifications of practices that can be 

implemented in new standards; 
123

• Monitoring should be at appropriate intensities (time and space) to 

determine effects of practices on water quality and quantity; 

• Monitoring should also measure co-benefits, including carbon 

sequestration, wildlife habitat, flood control, etc.; 

• Practices should be applied in suites or systems, and components should 

be monitored to determine effects of component practices; 

• Changes in hydrology and crop productivity must be measured in addition 

to changes in water quality. Even at the small-scale, too many studies 

have focused just on nutrient concentrations in outflow water and 

neglected hydrologic or productivity changes; 

• All components of the cost of adopting and maintaining these practices 

should be measured and monitored. Such costs include direct equipment 

and structural costs, yield effects, changes in management time, changes 

in risk, and other costs; 

• These studies should be designed to improve our understanding at local, 

medium and broad basin scale. Thus the experiments should be designed 

so that they can feed into conceptual models that operate at different 

scales; and 

• Within practical limits, studies should be part of an adaptive management 

research strategy for the MARB to optimize the efficiency of research 

investments and to assure that results are coordinated, complimentary and 

consistent. 

 Integrated modeling and monitoring play an important role in adaptive 

management. The cornerstone of adaptive management is the concept of learning about 

the impacts of actions and using that new understanding to guide future actions. Models 

can assist that learning by being used to evaluate impacts and uncertainties of proposed 

actions, such as targeted practices and locations or proposed policies, on both MARB and 

NGOM responses. In addition, monitoring must also be part of an adaptive management 

strategy in order to verify that the actions are addressing the stated goals or to test 

hypotheses. Monitoring is needed to improve the next generation of models and model 

assessments and to eventually verify that projected changes occur. 

 Adaptive management is also important to building infrastructure and to strategic 

planning and policy development of mechanisms of conservation practice 

implementation. For example, adaptive management can be used to evaluate if incentivebased programs are effective at bringing about changes in conservation practice 

acceptance and adoption at a local or small watershed level. At a basin level, other 

programs might be needed to facilitate adaptation of strategies and policies, and there 

must be constant feedback among all vested parties. As the scale of system increases 
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(i.e., from a small watershed to the entire MARB), the complexity of adaptive 

management increases dramatically. 

Key Findings and Recommendations 

 Adaptive management can be used at several scales of resolution in the NGOM 

and MARB to provide a framework under which management activities can occur while 

monitoring and modeling the outcomes in order to provide information so subsequent 

management can be improved. Therefore, the SAB Panel offers these recommendations. 

• An adaptive management approach should be adopted to evaluate the 

success of reaching goals and for testing hypotheses (at the relevant scale). 

• Conceptual models should be developed at appropriate scales of resolution 

to frame the adaptive management process in addressing factors affecting 

hypoxia in the NGOM. 

• Both the use of quantitative models and the collection of data should be 

conducted within an adaptive management framework and at appropriate 

management scales so that the information gained from models and data 

are related to the critical questions about managing and understanding the 

system. 

• Management actions should be designed as experiments within the context 

of evolving conceptual understanding of the system. The repercussions of 

management actions need to be monitored so the outcomes can be used to 

enhance learning and thus to improve future management actions. 

 
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4.2. Setting Targets for Nitrogen and Phosphorus Reduction 

 To reduce hypoxia in the bottom waters of the NGOM, the Integrated Assessment

set a target that N loading should be reduced by 30% in order to shrink the five-year 

running average size of the hypoxic zone to below 5,000 km2

 (1,930 mi2) by 2015. This 

reduction is significantly less than the three- to five-fold increase in N loading to the Gulf 

of Mexico due to human activity during the 20th century, and particularly in the last 30- 

to-50 years (Goolsby et al., 2001; Boyer and Howarth, in press). Since the Integrated 

Assessment, a number of modeling efforts have provided a better depiction of how the 

area of hypoxia may respond to reduced N loading. The three available models were 

compared by Scavia et al. (2004), who concluded from these models that the 30% 

reduction in N is probably not sufficient to reach the goal of a hypoxia area of 5,000 km2

or less (Scavia et al., 2004). The consensus from these models is that N loads probably 

need to be reduced by 40 to 45% to reach the hypoxia reduction goal. In addition, a 

number of studies suggest that the consequences of climate change need to be considered, 

and this may require an N load reduction on the order of 50 to 60% to meet the original 

Integrated Assessment goal for hypoxic area (Justic et al., 2003; Donner and Scavia, 

2007). However, predicting the consequences of climate change on nutrient fluxes and 

hypoxia remains a very uncertain business (Howarth et al., 2006). The SAB Panel finds 

that the consensus of models reported by Scavia et al. (2004) and the new model of 

Scavia and Donnelly (in press), which uses the latest available load estimates from the 

USGS, supports a target of reducing the five-year running average of N loadings by at 

least 45%. This target should be re-assessed as more monitoring data are obtained, 

current models are refined, and new models are developed. 

 Only recently has new evidence emerged for the need to control P inputs as well 

as N in the NGOM. Work by Sylvan et al. (2006) has shown P to be the limiting nutrient 

during periods of maximum primary production in the near-shore NGOM high 

productivity zone. Because previous attention has focused on N, there has been limited 

effort to model the effects of P on hypoxic area. Scavia and Donnelly (in press) used the 

previously developed and calibrated model (Scavia et al., 2004) to evaluate both the 

effects of new USGS load estimates and to assess the potential for P to control hypoxia 

dynamics under current and historical conditions. Confirming the results of Sylvan et al. 

(2006), Scavia and Donnelly found that P could have become limiting in some areas and 

times because of the relative increase in N loads during the 1970s and 1980s. While they 

concluded that P did frequently control hypoxia in near field zone of NGOM, they noted 

that a P only strategy would likely reduce production in the near field but possibly 

increase production in down-field N controlled areas of NGOM. Their work, using the 

new USGS load estimates, reinforced the need for a dual nutrient strategy combining a 

45% reduction in N with a 40 to 50% reduction in the five-year running average of P 

loading. While the far field effects could possibly be reduced through an N only strategy, 

they suggested that a prudent approach would be to reduce both N and P, simultaneously. 

They also noted that an N and P reduction strategy would not only reduce hypoxia in the 

NGOM but would also help to remove P-induced Clean Water Act impairments in the 

MARB. Based on this recent modeling work, the SAB Panel finds that a comparable P 
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reduction is needed, again based on 5-year running average fluxes. As with the N target, 

this P target should be re-assessed over time as more monitoring information is gained 

and new models are developed. 

Baseline for Reductions 

The CENR report and Scavia et al. (2004) made recommendations on an N 

reduction target with reference to average fluxes for 1980 to 1996. These fluxes were 

calculated using different methods (see Section 3.1) than in this report, but the N 

reduction target proposed recently by Scavia and Donnelly (in press) used a combination 

of the newer USGS five-yr LOADEST and composite estimates since 1980. In this 

report we only use the five-yr LOADEST results, since the composite estimates are 

incomplete; however, they are very similar to each other (again, see Section 3.1). 

During the last five years of record, annual water flux to the NGOM has declined 

by 5.8%, whereas nitrate-N and TKN have declined even more, leading to a total annual 

N reduction of about 21% (Table 9). Considering the original reduction target of a 30% 

reduction in total N, it would seem that substantial progress was made beyond the 

reduction that would occur from less flow alone. However, the largest reduction was in 

TKN, with a large part of this decrease from the Missouri River (discussed in Section 

3.1). For the important spring flux of N, there was little reduction in nitrate-N beyond the 

reduced water flow (-11 and -12.4 % declines in water and nitrate-N flux, respectively). 

Again, TKN was greatly reduced (-31.5%) during spring flows, leading to most of the 

decline in total N (-19.2%), beyond the reduction in water flux. This suggests that during 

the important high flow spring period (April, May, June), reductions in nitrate-N flux to 

the NGOM have not occurred under management systems and programs now in place 

since the last report. However, the annual nitrate-N reduction indicates that the tiledrained corn and soybean systems in the Upper Mississippi and Ohio River subbasins 

seem responsive on an annual basis to the recent reductions in net N inputs, as discussed 

in Section 3.2. Whether spring nitrate-N loads will respond to these changes in NANI is 

uncertain at this time. 
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Table 9: Annual and spring (sum of April, May, June) average flow and N and P fluxes for the MARB for 

the 1980 to 1996 reference period compared to the most recent five year period (2001 to 2005). Load 

reductions in mass of N or P also shown. 

 1980 to 

1996 flux 

2001 to 2005 

flux 

change 45% 

reduction N 

target flux 

45% reduction 

P target flux 

 million m3 (water) or million 

metric tons 

% million metric tons 

Annual 

 Water 692,500 652,500 -5.8 

 Nitrate-N 0.96 0.81 -15.4 0.53 

 TKN 0.61 0.43 -30.0 0.34 

 Total N 1.58 1.24 -21.1 0.87 

 Total P 0.137 0.154 +12.2 0.075 

 

Spring 

 Water 236,800 210,600 -11.0 

 Nitrate-N 0.38 0.33 -12.4 0.21 

 TKN 0.21 0.14 -31.5 0.12 

 Total N 0.59 0.48 -19.2 0.32 

 Total P 0.046 0.050 +9.5 0.025 

 For total P flux, both annually and during the spring, there were increases of 12.2 

and 9.5%, respectively. It is not clear why total P fluxes are increasing (with 

corresponding smaller water fluxes), and the result suggests that the reduction target of 

45%, relative to the 1980 to 1996 period, is close to 50% for the 2001 to 2005 period. 

Likewise, the 45% N load reduction target, relative to the 1980 to 1996 period, is 

equivalent to a 30% reduction relative to the 2001 to 2005 period. Fertilizer P 

consumption in the MARB has been relatively constant since about 1984 and is similar to 

consumption during 1970-to-1975 period. Net P inputs to the MARB have declined since 

the 1970s and have been predominantly negative since the mid-1990s (see Section 3.2 

and Figure 34). Table 9 also indicates N and P reduction recommendations in units of 

mass with reduction targets of 45% N and 45% P, assuming the reduction were spread 

across all forms of N and P, that occur both annually and during the spring. 

While the SAB Panel finds that both N and P reductions are warranted, additional 

modeling and dose response research is needed to refine the reduction targets, 

particularly for P loading. Scavia and Donnelly (in press) presented the only model 

results that relate P loads to hypoxia in the NGOM. Further, there are no experimental 

data relating phytoplankton responses there to different levels of P. Ideally, targets for 

reducing P based on water quality should have greater model support, and should 

consider dose response relationships for P responses by the in situ phytoplankton 

communities. In the meantime, the response of the Gulf system to a specific amount of P 
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reduction remains uncertain and must await the formulation of new models and dose 

response relationships for the receiving waters. Water quality models aimed at 

evaluating the effects of these reductions will also rely on this information. Dose 

response relationships should be developed using in situ bioassays designed to “ask the 

phytoplankton” what the response relationships and bloom thresholds are. These 

bioassay experiments are a logical follow-up to the work of Sylvan et al. (2006), which 

has shown P to be the limiting nutrient during periods of maximum primary production in 

the near-shore NGOM high productivity zone. Bioassays are needed on a seasonal basis, 

where the effects of hydrologic variability and changing N:P input (loading) ratios on 

primary production, phytoplankton community composition, and biogeochemical and 

trophic fate can be evaluated. 

 In Section 4.5.8 on Most Effective Actions for Industrial and Municipal Sources, 

the SAB Panel provides some ballpark estimates of possible N and P reductions from 

upgrading major municipal wastewater treatment plants. The SAB Panel’s example 

calculations demonstrate that sewage treatment plant upgrades to achieve total N 

concentration limits of 3 mg/L and total P concentrations of 0.3 mg/L could create 

reductions in total annual N flux to the Gulf by about 10% and the total spring N flux by 

about 6%. Upgrading to achieve P concentrations of 0.3 mg/L would create reductions in 

P fluxes from sewage treatment plants from 41,000 metric tons P/yr (45,000 ton P/yr) to 

10,500 metric tons P/yr (11,600 ton P.yr) or about a 75% reduction in annual flux from 

sewage treatment plants to the MARB. These reductions, in turn, would translate into 

reductions of total annual P flux to the Gulf by about 20% and the total spring P flux by 

about 15%. If further investigation and data collection confirms the SAB Panel’s 

calculations, upgrades to major wastewater treatment plants in the MARB could 

accomplish nearly half of the Panel’s recommended P reduction targets. This would 

represent very significant progress for both improving water quality in the MARB and 

reducing hypoxia in the NGOM. 

 Despite the need for additional model and bioassay work, the proposed target of a 

45% reduction in annual P load should be used in an adaptive management framework to 

allow development of strategies that optimize both N and P reductions while more 

knowledge is acquired on P reduction impacts on near-field hypoxia. Unlike N, the P 

reduction strategy will help address water quality impairments in the MARB. Given the 

evidence that both N and P should be reduced in the NGOM, setting a goal for P 

reduction should not await the development of new models and availability of new 

experimental data. Enough information exists now to set a goal in an adaptive 

management context beginning with the P reductions that are already feasible given 

existing technologies and options. 

 In 2000, EPA recommended nutrient criteria to States and Tribes for use in 

establishing their water quality standards consistent with Section 303(c) of the Clean 

Water Act (CWA) (USEPA, 2000). EPA’s recommended criteria represent an estimated 

“reference condition,” and it is assumed that the reference condition concentration would 

protect all designated uses (including the most protected uses, such as high quality 

fisheries, sensitive aquatic life, etc.). The SAB Panel asked EPA for a comparison of the 
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SAB Panel’s recommended 45% reductions for TN and TP flux to the reductions in 

nutrient levels that would correspond to EPA’s ecoregional nutrient criteria for reference 

conditions (U.S. EPA, 2006b). This comparison is provided in Appendix C: EPA’s 

Guidance on Nutrient Criteria. Although a number of assumptions were required to 

make this comparison (see the caveats in Appendix C), EPA's preliminary analysis 

suggests that the SAB Panel's recommended targets for reducing TN and TP are, for most 

regions, not likely to be as stringent as would be obtained if states adopted EPA’s 

recommended reference condition values into state water quality standards for all waters. 

This comparison should not be interpreted as the SAB Panel’s endorsement of EPA’s 

recommended nutrient criteria but rather an emphasis on the need to consider both within 

basin nutrient criteria and NGOM load reduction goals. Numeric nutrient standards being 

developed by the states of the MARB will almost certainly be concentration rather than 

load based and may be most stringent during warmer, lower flow periods when absolute 

loads can be relatively low but when local waters are most frequently impaired by excess 

nutrient levels. It will be important for EPA and other agencies to evaluate and, if 

necessary, reconcile within-basin water-quality standards with load-reduction goals for 

the NGOM. Strategies are needed for integrating standards throughout the MARB to 

better manage hypoxia as well as local water quality. 

A mechanism in the Clean Water Act for addressing water quality impairments is 

the development of Total Maximum Daily Loads (TMDLs), though it is important to note 

that the focus of TMDL development is identification of the source and causes of water 

quality impairment, rather than on implementation of change for improving water quality. 

Under Section 303(d) of the Clean Water Act, states, territories, and authorized tribes are 

required to develop lists of impaired waters (i.e., waters that have not met water quality 

standards). The law requires that the appropriate jurisdictions develop TMDLs for these 

impaired waters. The TMDLs specify the maximum amounts of pollutants that 

waterbodies can receive and still meet water quality standards. In addition, TMDLs 

allocate pollutant loadings among point and non-point sources. 

The status of nutrient criteria and TMDL development along the Mississippi 

River has been reviewed by the National Academy of Sciences (National Academy of 

Sciences, 2007). The National Academy of Sciences notes that none of the 10 

Mississippi River mainstem states currently have numeric criteria for nitrogen or 

phosphorus applicable to the River, and that without such standards, there is little 

prospect of significantly reducing or eliminating hypoxia in the Gulf of Mexico. The 

National Academy of Sciences also describes how the process of developing numeric 

nutrient criteria and TMDLs for the Mississippi River could lead to water quality 

improvements in the Gulf of Mexico. NAS suggests that through such a process, EPA 

could adopt the necessary numerical nutrient criteria for the terminus of the Mississippi 

River and waters of the northern Gulf of Mexico. Maximum nutrient loads could be 

assigned to each state and the loads could be translated into water quality criteria. Each 

state would then be required to develop a TMDL for waters that failed to meet the 

applicable criteria, and a coordinated effort could be undertaken to reduce point and nonpoint source loads to meet allocations established by the TMDLs. Thus, the NAS report 

identifies an approach through existing legislation (the Clean Water Act) that could be 
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used to redress Gulf Hypoxia, but the SAB stresses that a great many steps exist between 

calling for “a coordinated effort” and implementing the full set of actions that must be 

undertaken for water quality to actually improve in the Gulf. 

Key Findings and Recommendations 

Based on findings since the Integrated Assessment, a N reduction target of 

greater than 30% will be needed to reduce the hypoxic area to 5000 km2

 (1,930 mi2). 

Recent research indicates N reductions of at least 45% will be needed to achieve the 

target in most years and reductions may have to exceed 50% due to effects of climate 

change. Research by several investigators provides evidence that P may limit primary 

production in the river outflow, near-field areas of the Gulf. Based on new research 

with the same model used to establish the N target, reductions in P loads of 40 - 50 % 

are needed to reduce P-controlled hypoxia in the near-field areas of NGOM. P 

reductions in the MARB will not only benefit the NGOM but will also help to address P 

impairments in the MARB. Based on these findings, the SAB Panel offers the following 

recommendations. 

• To reduce the size of the hypoxic zone, the total N flux to the NGOM from the 

combined Mississippi and Atchafalaya Rivers must be reduced by at least 45% 

from 1980 to 1996 average fluxes, to no more than 790,000 metric tonne N/y 

(870,000 ton/yr), and 290,000 metric tonne N (320,000 ton) during the spring 

(April, May, June), both on a five-year running average. 

• To reduce the size of the hypoxic zone, commensurate reductions in P are 

needed. The total P flux to the NGOM from the combined Mississippi and 

Atchafalaya Rivers should be reduced by at least 45% from 1980 to 1996 

average fluxes, to no more than 68,000 metric tonne P/yr (75,000 ton P/yr) on a 

five-year running average. 
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4.3. Protecting Water Quality and Social Welfare in the Basin 

 The SAB Panel has been asked whether social welfare can be protected while 

reducing hypoxia and improving water quality in the Basin. To thoroughly answer this 

question would require quantification of the full costs of all activities undertaken to 

reduce the necessary nutrient loading into the Gulf (from agricultural sources, point 

sources, air deposition, etc.) and the full benefits accruing from those activities. The 

benefits would include the direct benefits of reducing the size of the hypoxic zone 

(commercial fishery effects, recreational fishery gains, the value placed on preserving 

intact ecosystems, biodiversity, etc.) and the “co-benefits” (such as improved local water 

quality, increased wildlife habitat, flood control, aesthetic values, etc.). 

Since the costs, benefits, and co-benefits will depend on the extent of coverage 

and specific locations that control options are located, a complete answer to the question 

would require knowing the details of how such nutrient reductions would occur. For 

example, if these reductions are to be achieved entirely through restoration of wetlands 

and tighter municipal source controls, it would be necessary to know where the wetlands 

would be located and where the point source reductions would occur in order to estimate 

their costs and their co-benefits. In contrast, an entirely different set of co-benefits and 

costs would likely result from relying on a broader array of control options that also 

included nutrient management, increased perennials, riparian buffers, drainage 

management, and reductions in air deposition. Further, the exact policy approach (e.g., 

expanded EQIP funding, mandates, or taxes) would need to be specified if estimates of 

the incidence of the costs are to be estimated (i.e., whether the costs would ultimately be 

borne by taxpayers, consumers, or by farmers and landowners). 

To date, no set of models and/or studies have been undertaken that address all of 

the necessary components on a basin-wide scale to estimate the effects on social welfare. 

However, a number of studies, beginning with the research in the Integrated Assessment, 

have been done that address substantial components of this question. More complete 

efforts at quantifying the control costs than the benefits have been undertaken, though 

there remains a need for much more work on both sides of the equation. Integrated 

models at multiple levels and scales are needed to support this effort. The existing 

research focuses largely on agricultural non-point source control. This section 

summarizes findings from the limited set of large-scale economic-watershed models of 

agricultural non-point sources that have been applied to date. 

Assessment and review of the cost estimates from the CENR Integrated Assessment 

 Doering et al. (1999) in the Integrated Assessment undertook an ambitious costeffectiveness analysis of several policy approaches to reach the N loss reduction goal of 

20% established as part of the Integrated Assessment. The central modeling system they 

used was the U.S. Mathematical Programming (USMP) model, which represents the 

agricultural sector in 45 production regions throughout the United States with 10 crops, 

16 animal products, retail and processed products, and a range of domestic and 
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international supply and demand relationships. Management practices include crop 

rotations, five tillage options, and varying fertilizer rates. 

 The environmental effects of various management practices and land uses in 

USMP are predicted by the EPIC model (the Environment Productivity Impact 

Calculator). USMP uses EPIC to predict changes in N loss, P loss, and sediment loss at 

the edge of the field from changes in land use and conservation practices. Donner et al. 

(1999) chose a 20% N loss reduction goal as “the best combination of sizable nitrogenloss reductions and acceptable economic costs” (Doering et al., (1999) page 37). The 

remainder of their analyses focused on the evaluation of several policies that might 

achieve this environmental goal. Some key predictions from the modeling system 

include: 

• A 20% reduction in fertilizer N application rates would result in the reduction 

of edge-of-field N loss by about 11%. In contrast, a 45% reduction mandate 

and fertilizer tax set to achieve a 45% reduction is predicted to result in the 

target goal of N loss reduction of about 20%. The less than proportional 

reduction in N loss coming from reduced fertilization in this modeling system 

is a result of predicted changes in acreage resulting from the feedback effect 

of price changes. Specifically, higher crop prices due to lower yields from the 

reduced fertilization rates induce more acreage planted to the fertilized crop, 

thereby partially offsetting the reduction in N. Whether the magnitude of the 

yield effects embedded in these models is accurate is an important question. 

For further discussion of this issue, see Section 4.5.6. 

• Some 7.29 million hectares (18 million acres) of wetland restoration would 

achieve the 20% reduction in N loss goal at a cost of over $30 billion. 

• Restoration of 10.9 million hectares (27 million acres) of riparian buffers was 

estimated to cost over $40 billion and generated relatively small reductions in 

N losses, suggesting that this strategy is not cost-effective for hypoxic zone 

control. In light of current evidence that phosphorous is also of concern, this 

result should be reconsidered as there is significant evidence that buffers can 

be quite effective in holding sediment and phosphorous in field. 

• A “mixed policy” with a 2.02 million hectares (5 million acre) wetland 

restoration program in conjunction with a 20% fertilizer reduction is more 

cost-effective than most of the previous approaches, but the 45% reduction in 

fertilizer is more cost-effective yet. 

• The introduction of point-non-point source trading across the basin where the 

cap applies only to point sources will not achieve the 20% N loss reduction 

due to the relatively small magnitude of N contribution from point sources. 

Even with a stringent standard on point sources, only about 5% of the needed 

reductions occur. 
133

• These policies are likely to produce large “co-benefits” (i.e., other 

environmental benefits occurring within the basin and on-farm productivity 

benefits not immediately captured in the current profitability resulting from 

the policies). For example, the authors estimate that restoration of 405,000 

hectares (1,000,000 acres) of wetlands would yield total benefits in the basin 

that exceed the costs, even without considering any benefits of hypoxia 

reduction. 

Cost estimates used for the Integrated Assessment for a 20% reduction in N 

discharge coming from agricultural non-point sources range from $15 billion to $30 

billion; however these estimates suffer from a number of shortcomings including 

consideration of only a few options for reducing nutrient discharge and limited targeting. 

More inclusive assessments with better targeting of options to locations where they are 

most appropriate may reduce these costs. 

In follow up research, some of the same study coauthors (Ribaudo et al., 2001) 

compare nitrogen reduction methods with wetland restoration and low and high levels of 

N loss reduction. They find that nutrient management is more cost-effective at low levels 

of N loss reduction while wetlands restoration is more cost-effective at high levels. Table 

10 and Table 11 (listed at the end of this discussion) briefly summarize the key 

components of these studies and the other large-scale studies that are reviewed in the 

following discussion. 

 Due to limits on the understanding of the economics and natural science at the 

time, the work in the Integrated Assessment and its follow up is based on assumptions 

that, in light of more recent research and availability of data, could be improved upon in 

future work. The USMP model represents a wide variety of agricultural raw inputs and 

intermediate products at a relatively aggregate scale. However it does not contain 

detailed description of land use, soil characteristics, yields, etc. at the individual field 

and/or sub basin scale. This inability to target finer scales could result in overstating the 

costs of meeting a particular reduction goal because significant cost savings can accrue 

from targeting land-management strategies. 

 The Integrated Assessment assumed a one-to-one relationship between the 

reduction in edge-of-field nitrogen loss and reduced loadings to waterways without 

incorporating the geographic differences in movement of N from the field of origination 

to the Gulf. Whether this shortcoming over- or under-states the costs is an empirical 

question, but the results coming from a model that explicitly incorporates the fate and 

transport of nutrients and sediment might suggest very different results concerning the 

cost-effectiveness. 

Other large scale integrated economic and biophysical models for agricultural non-point 

sources 

 Since completion of the Integrated Assessment, several basin-wide studies have 

evaluated policies that might reduce Gulf hypoxia and/or have effects on other 
134

environmental amenities that could be considered co-benefits (including carbon 

sequestration and upstream, local water quality indicators). The models can be divided 

into those that use the USMP modeling framework and those based on econometric 

estimates of behavioral response to economic drivers. 

 Booth and Campbell (2007) used a regression model to estimate the cost of 

reducing N losses when targeting conservation dollars to those areas with the highest 

proportion of fertilizer use. They modeled a hypothetical case in which conservation 

enrollment rises in direct proportion to the nonlinear rise in nitrate flux that occurs as 

fertilization intensity increases. The result was an increase in the amount of land in the 

high-fertilizer watersheds enrolled in the Conservation Reserve Program by 2.7 million 

hectares (6.67 million acres) (a 29% increase over 2003 CRP levels) at a cost of $448 

million. Booth and Campbell (2007) describe this as a 6.2% increase over the combined 

cost of commodity support and conservation programs. They account for the drop in 

commodity support spending that would accompany the enrollment of commodityfarmed land in the CRP. Booth and Campbell (2007) do not specify the percentage 

reduction in nitrate loading that would result from this scenario. 

Wu et al. (2004) and Wu and Tanaka (2005) developed an econometric model of 

crop choice and tillage choice using the National Resources Inventory for the upper 

Mississippi River basin. They estimated the probability of adopting conservation tillage 

and crop choice based on a variety of physical and economic variables including land 

quality, slope, climate conditions, and profits. They used over 40,000 crop land points 

observed for 16 years, although only a subset of the observations were used for model 

fitting. These adoption models then simulate adoption profiles under alternative policies. 

Finally, the environmental effects of the policies are predicted with a biophysical model. 

Wu et al. (2004) used a set of environmental production functions estimated via a metamodeling approach (Wu and Babcock, 1999), based on data generated from the EPIC 

model. They found that crop rotations are not a cost-effective strategy to N reduction. 

Wu and Tanaka (2005) used the SWAT model to predict water quality changes 

from the policies. They considered the same two policies as Wu et al. (2004), as well as 

a policy that would increase the amount of land set-aside in a Conservation Reserve-type 

program and a fertilizer tax at various rates. They found a fertilizer tax to be the most 

cost-effective of policies they considered. 

Kling et al. (2006) employed a similar econometric modeling approach. Like Wu 

et al. (2004), they used the National Resource Inventory data to link the cost data with the 

SWAT model. They estimated the costs and water quality benefits of implementing a set 

of conservation practices associated with implementation rules based on distances to a 

waterway, slope, and erodibility indices. The conservation practices assessed include 

grassed waterways, nitrogen management, terraces, buffers, land retirement and 

conservation tillage. They estimated that this placement of conservation practices on the 

landscape would cost over $800 million annually (or roughly $16 billion if viewed as a 

lump sum cost assuming a 5% rate of discount) and would achieve a 22% reduction in N 

loadings into the upper Mississippi River basin at Grafton Ill. Within the UMRB, they 
135

estimated a 40-66% reduction in sediment loads, a 6-47% reduction in P loads, and a 9-

29% reduction in N loads. These estimates (like those from all of the studies reviewed 

here) are likely to be very sensitive to the set of conservation practices included and the 

specific scenarios studied. 

Greenhalgh and Sauer (2003) used the USMP, augmented in two important ways: 

1) they configured the model by watersheds and added information on municipal waste 

water treatment plants and 2) they included “attenuation” coefficients derived from the 

SPARROW model to reflect the transport component of N flows between watersheds. 

The focus of their work was on policy options for hypoxia that also contribute to 

greenhouse gas reductions. The policies they considered include N trading between point 

and non-point sources, GHG trading assuming external carbon prices of $5/ton and 

$14/ton, N trading with additional payments for GHG emission reductions, an N fertilizer 

tax, a subsidy to farmers willing to shift from conventional to conservation tillage, and an 

expansion of the CRP program to 16 million ha (40 million ac) nationwide. Of the 

policies evaluated, none achieved the 20% reduction goal of the Doering et al. (1999) 

analysis. The largest reductions were achieved in their simulation of point/non-point 

source trading with a stringent N standard. The most cost-effective policies were also the 

trading programs. 

Ribaudo et al. (2005) also considered the possibility of N trading between point 

and non-point sources using the USMP model. They found that trading has significant 

potential to reduce costs relative to a requirement that wastewater treatment plants be 

required to install stringent nutrient removal technology.2

 These studies shed light on the costs of addressing the hypoxia problem from 

conservation practices in the agricultural sector, and the way these costs may vary 

depending on the policy instrument chosen (trading program, conservation payment, tax, 

etc.). These studies also directly bear on the question of how much it will cost to address 

local water quality in the MARB. However, as noted above, shortcomings of the 

integrated models have prevented assessment of many policies as well as conservation 

practices and sinks. None of the models include point source and non-point source 

control options. With the exception of Booth and Campbell (2007), most models have 

not adequately addressed the cost savings associated with targeting. Nonetheless, results 

to date suggest that there is large variability in the costs of alternative policies. The issue 

of who pays these costs may also be important to consider since the incidence (who must 

pay the costs) may differ dramatically across policies. A notable example is a fertilizer 

tax, which has the same social costs as a restriction but which may have a much higher 

incidence on farmers. 

Improved estimates of the costs of installing and maintaining conservation 

practices could be generated with the current suite of models by considering alternative 

sets of conservation practices. This can be accomplished using the following steps: 1) 

identifying conservation practices that are most likely to be effective in reducing nutrients 

 

2

 It is important to recognize that these studies assume a perfectly efficient water quality trading program 

with no trading restrictions; current water quality trading programs do not match the modeled system. 
136

important for hypoxia, and 2) identifying scenarios that place these conservation practices 

on the landscape. These scenarios could be based on rules of thumb (identifying for 

example a particular conservation practice to be used on cropland with specific climate 

and soil characteristics), algorithms for optimal placement to minimize costs, multiple 

goals such as maximizing in basin co-benefits or income support, or policy relevant 

methods such as the use of an environmental benefits index, etc.; and 3) computing cost 

estimates from economic models and water quality changes from watershed models. 

Research Assessing the Basinwide Co-Benefits 

 As noted above, many of the same practices that could contribute to reductions in 

the hypoxic zone could also have significant effects on local water quality, carbon 

sequestration, wildlife habitat, flood control, and other ecosystem services. The physical 

co-benefits of many conservation practices and sinks are described in Section 4.5.10. On 

the basin-wide scale, there are a few studies that provide physical measures of one or 

more co-benefits that are associated with implementation of conservation practices that 

would address hypoxia, particularly related to carbon sequestration and water quality (see 

for example, Feng, et al. (2005), Lewandrowski et al. (2004), Greenhalgh and Sauer 

(2003)). These studies consistently indicated that significant co-benefits are present, but 

these estimates are not monetized and are reported in physical units. Further, the policies 

analyzed are not focused on hypoxia reduction. 

Thus, the work reported in the Integrated Assessment remains the most complete 

coverage to date of the potential value to MARB residents of the water quality and other 

co-benefits. The estimates provided there suggested that the monetized value of the 

benefits to the basin were larger than the costs based primarily on benefit estimates of the 

value of erosion control and wetlands restoration. A more complete accounting of these 

benefits could be developed using benefits transfer techniques, although there are many 

ecosystem services for which currently accepted methods are not likely to adequately 

fully capture the value of the benefits. But, in any case, because the Integrated 

Assessment was not able to quantify all co-benefits, total co-benefits within the basin 

would almost certainly be larger than those estimated. 

Due to the incredible complexity in this system, as well as limits in data, 

modeling and research, definitive statements on social welfare are not possible. For 

example, there is incomplete information on the costs of farm-level actions to reduce 

edge-of-field nutrient losses. There is even greater uncertainty in quantifying the 

effectiveness of farm-level nutrient control actions in reducing watershed-level nutrient 

flux and about the relationship between watershed-level nutrient flux and the spatial and 

temporal dimensions of the hypoxic zone. These uncertainties are further exacerbated by 

the possibility of regime shift in the Gulf of Mexico, whereby the system could become 

more susceptible to hypoxia following the initial occurrences. If regime shift is a factor, 

then historic data on the relationship between nutrient flux and the size of the hypoxic 

zone does not provide guidance on the decrease in nutrients required to achieve a given 

reduction in the size of the hypoxic zone. Hence, a return to historic lower levels of 
137

nutrient fluxes might not be adequate to return to a corresponding size of the hypoxic 

zone. 

 There are many sources of uncertainty in the economic, hydrologic, and Gulf 

systems that make it difficult to render definitive conclusions about social welfare 

Indeed, it is precisely because of these many uncertainties and need for additional 

research that we recommend an approach based on an adaptive management strategy that 

aims to move in a “directionally correct” fashion, rather focusing on achieving a precise 

outcome. 

While we cannot definitely say that we can achieve the 5,000 km2

 (1,930 mi2) 

goal while maintaining social welfare, there is evidence that suggests it is feasible to do 

so. First, and perhaps most importantly, welfare losses in the Basin will be at least 

partially or even totally offset by co-benefits of nutrient reduction actions. For example, 

if wetlands restoration is used to control nutrient flux, it will result in improvements in 

wildlife habitat and local water quality, both of which will improve welfare in the Basin. 

Findings from the Doering et al. (1999) assessment point out that the benefits accruing 

locally from wetlands restoration might well exceed the costs, even without any Gulf 

hypoxia reductions. Similar estimates are reported in Hey et al. (2004) for substantial 

restoration of wetlands in flood plains (see Section 4.4.2). Management actions that 

reduce farm-level nutrient losses may lead to better local water quality, thereby 

improving welfare for affected residents within the Basin. If management actions are 

undertaken to control air emissions, thereby reducing atmospheric deposition of nitrogen, 

it will result in improvements in air quality, reduction in acid precipitation, lower 

emissions of greenhouse gasses, etc. Thus, co-benefits within the Basin will at least 

partially and perhaps fully offset welfare losses associated with the costs of implementing 

management actions. And in the longer term, a transition from corn to perennial crops 

could benefit farmers and other Basin residents. Thus, there may be larger scale 

transitions in the agronomic system that provides opportunities to reduce nutrient flux 

while maintaining welfare in the Basin. 

A second reason for optimism is that cost-effective approaches, such as targeting 

low cost sources and using emissions trading, have not yet been applied. These 

approaches have the potential to reduce the costs of nutrient control, possibly 

considerably, thereby reducing the burden of complying with the goal. Thus, there may 

be opportunities to control the cost of nutrient reduction. 
138

Table 10: Summary of Study features of Basin wide Integrated Economic-Biophysical Models 

Authors Study Region Models used Environmental 

measures 

Comments 

Doering et al. 

(1999) 

Entire U.S. 

with policies 

simulated in 

Mississippi 

River basin 

USMP and EPIC N, P, and sediment in 

the MARB ( not 

delivered to the 

NGOM) 

Original CENR 

study 

Ribaudo et al. 

(2001) 

Entire U.S. 

with policies 

simulated in 

Mississippi 

River basin 

USMP and EPIC N, P, and sediment in 

the MARB ( not 

delivered to the 

NGOM) 

Extension of 

CENR study 

Greenhalgh and 

Sauer (2003) 

Entire U.S. 

with policies 

simulated in 

Mississippi 

River basin 

USMP and EPIC 

with Sparrow 

derived transport 

coefficients 

N delivered to the 

Gulf, greenhouse gas 

emissions, P and N, 

soil erosion in the 

MARB 

Study focuses on 

co-benefits of 

policies 

Wu et al. (2004) upper 

Mississippi 

River basin 

Econometric model 

and EPIC based 

metamodels 

N leaching, N runoff, 

wind erosion, and 

water erosion in 

UMRB 

Finer spatial detail 

than USMP but no 

price feedbacks 

Ribaudo et al. 

(2005) 

Entire U.S. 

with policies 

simulated in 

Mississippi 

River basin 

USMP and EPIC N in MARB Follow up to 

original CENR 

study 

Wu and Tanaka 

(2005) 

upper 

Mississippi 

River basin 

Econometric model 

and SWAT 

N delivered to the 

NGOM 

Finer spatial detail 

than USMP but no 

price feedbacks 

Kling et al. (2006) upper 

Mississippi 

River basin 

Econometric model 

and SWAT 

N, P, and sediment in 

UMRB and N 

delivery to the 

NGOM 

Finer spatial, but 

no price feedbacks 
139

Table 11: Summary of Policies and Findings from Integrated Economic-Biophysical Models 

Study Policies/Actions Evaluated Key Findingsa 

Doering et al. 

(1999) 

1. Fertilizer reduction 

mandates/fertilizer taxes 

2. Wetland restoration 

3. Riparian Buffer 

4. Mixed Policy (wetlands and 

fertilizer reduction) 

5. Water Quality Trading 

1. Cost effective approaches exist to reducing nitrogen 

losses in the 20% range 

2.Wetland-based strategies are more expensive than 

fertilizer reduction 

3. Buffers are not cost-effective for reducing N losses 

4. A combination of 5 million acre wetland restoration 

with 20% fertilizer reduction is most cost-effective 

5. These cost-effectiveness measures do not take into 

account the transport of nitrogen to the Gulf and the 

rankings of preferred alternatives could change 

Ribaudo et al. 

(2001) 

1. Reduce fertilizer rates 

2. Wetland restoration 

1. below 26% reduction in N losses, fertilizer 

reduction/management is most cost-effective 

2. Above this rate, wetland restoration is most costeffective 

Greenhalgh 

and Sauer 

(2003) 

1. N trading between point and nonpoint sources, 

2. greenhouse gas trading 

3. N trading with additional 

payments for GHG reduction 

4. N fertilizer tax, 

5. conservation tillage payment 

6. expansion of CRP to 40 million 

acres nationwide 

1. Nutrient trading (point/non point) with tighter 

discharge limits could reduce nitrogen reach the 

NGOM by 11% annually 

2. Nutrient and greenhouse gas trading were the lowest 

cost policies, but nutrient trading was the most costeffective 

3. The co-benefits of these policies in terms of 

greenhouse gas reductions, phosphorous, and sediment 

can be significant 

Wu et al. 

(2004) 

1. Conservation payments for 

conservation tillage and 

2crop rotations 

Crop rotations not a cost-effective strategy for N 

reduction 

Wu and 

Tanaka (2005) 

1. Fertilizer tax 

2. Payments for conservation tillage 

3. Payments for land retirement 

4. Payments for crop rotations 

Fertilizer tax is the most cost-effective of policies 

considered 

Booth and 

Campbell 

(2007) 

Targeting CRP to watersheds with 

the greater proportion of fertilizer 

used. Hence CRP rises in direct 

proportion to fertilizer/cropping 

intensity. 

Targeting CRP and enrolling an additional 2.7 million 

hectares in those areas with the greatest fertilizer 

intensity would increase annual agricultural subsidies to 

the MARB by 6.2% (over the combined commodity 

support and conservation funding in 2003). 

Ribaudo et al. 

(2005) 

N trading between point and nonpoint sources Trading between waste water treatment plants and nonpoint/agricultural sources to meet the reductions 

achievable by installing advance nutrient removal 

technology at treatment plants would have large welfare 

gains 

Kling et al. 

(2006) 

Implementation of a set of targeted 

conservation practices including 

conservation tillage, land 

retirement, terraces, contouring, 

grassed waterways, and reduce 

fertilization rate on corn 

1. Annual costs of $800 million per is predicted to 

achieve 22% reduction in N loading to the NGOM, 

2. within the UMRB sediments loads were reduced by 

40-66%, total P was reduced by 6-47% and N by 9-29% 

a) Doering et al. (1999) also conclude that fertilizer restrictions are more cost-effective than a fertilizer tax, but they 

apparently incorrectly count tax revenues as a cost rather than a transfer. The restrictions and tax have the same welfare 

effects, though different distributional implications. 
140

Principles of Landscape Design 

 Another perspective for protecting social welfare can be drawn from the 

principles of landscape design. A landscape perspective involves broad-scale 

consideration of how decisions affect resources, particularly in the long run. Guidelines 

have been proposed as a way to facilitate land managers considering the ecological 

ramifications of land-use decisions (Dale et al. 2000). These guidelines are meant to be 

flexible and to apply to diverse land-use situations, yet require that decisions be made 

within an appropriate spatial and temporal context. These landscape design guidelines 

can serve as a checklist of factors to be considered in making decisions that relate to 

implications for hypoxia in the Gulf. 

• Examine the impacts of local decisions in a regional context. The spatial array of 

habitats and ecosystems shapes local conditions and responses (e.g., Patterson, 1987; 

Risser, 1985), and local changes can have broad-scale impacts over the landscape. 

Hypoxia is a classic example of such impacts, for fertilizer applications in the 

Midwestern states can affect oxygen conditions in the Gulf of Mexico. This 

guideline notes that it is critical to examine both the constraints placed on a location 

by the regional conditions and the implications of decisions for the larger area. 

Therefore, it is critical to identify the surrounding region that is likely to affect and 

be affected by the decision and examine how adjoining jurisdictions are using and 

managing their lands. Forman (1995) suggests that land-use planning should first 

determine nature’s arrangement of landscape elements and land cover and then 

consider optimal spatial arrangements and existing human uses. Following this 

initial step, he suggests that the desired landscape mosaic be planned first for water 

and biodiversity; then for cultivation, grazing, and wood products; then for sewage 

and other wastes; and finally for homes and industry. Of course, planning under 

pristine conditions is typically not possible. Rather, the extant state of development 

of the region generally constrains opportunities for land management. 

• Plan for long-term change and unexpected events. Impacts of decisions can, and 

often do, vary over time as a result of delayed and cumulative effects. Future options 

are often constrained by the decisions made today as well as by those made in the 

past. For example, areas that are urbanized are unlikely to be available for any other 

land uses because urbanization locks in a pattern on the landscape that is hard to 

reverse. Thus, management actions should be implemented with some consideration 

as to the physical, biological, aesthetic or economic constraints that are placed on 

future uses of resources. External effects can extend beyond the boundaries of 

individual ownership and thus have the potential to affect surrounding owners. 

Planning for the long term also requires consideration of the potential for unexpected 

events, such as variations in temperature or precipitation patterns or disturbances. 

Long-term planning must also recognize that one cannot simply extrapolate 

historical land-use impacts forward to predict future consequences of land use. The 

transitions of land from one use or cover type to another often are not stable over 

time because of changes in demographics, public policy, market economies, and 

technological and ecological factors. 
141

• Preserve rare landscape elements, critical habitats, and associated species. This 

guideline implies a hierarchy of flexibility, and it implicitly recognizes ecological 

constraints as the primary determinants in this hierarchy. For example, a viable 

housing site is much more flexible in placement than an agricultural area or a 

wetland dedicated to improving water quality and sustaining wildlife. Optimizing 

concurrently for several objectives requires that planners recognize lower site 

flexibility of some uses than others. However, given that most situations involve 

existing land uses and built structures, this guideline calls for examining local 

decisions within the regional context of ecological concerns as well as in relation to 

the social, economic, and political perspectives that are typically considered. 

• Avoid land uses that deplete natural resources over a broad area. Depletion of 

natural resources disrupts natural processes in ways that often are irreversible over 

long periods of time. The loss of soil via erosion that can occur during agriculture 

and the loss of wetlands and their associated ecological processes and species are 

two examples. This guideline requires the determination of resources at risk, which 

is an ongoing process as the abundance and distribution of resources change. This 

guideline also calls for the deliberation of ways to avoid actions that would 

jeopardize natural resources and recognition that some land actions are inappropriate 

in a particular setting or time, and they should be avoided. 

• Avoid or compensate for effects of land use on ecological processes. Negative 

impacts of land use practices might be avoided or mitigated by some forethought. 

To do so, potential impacts need to be examined at the appropriate scale. At a fine 

scale, farm practices may interrupt ecoregional processes. At a broad scale, patterns 

of watershed processes may be altered, for example, by changing drainage patterns 

as part of the land use. Therefore, how proposed actions might affect other systems 

(or lands) should be examined. For example, human uses of the land should avoid 

uses that might have a negative impact on other systems; at the very least, ways to 

compensate for those anticipated effects should be determined. It is useful to look 

for opportunities to design land use to benefit or enhance the ecological attributes of 

a region. 

 

• Implement land-use and -management practices that are compatible with the 

natural potential of the area. Local physical and biotic conditions affect ecological 

processes. Therefore, the natural potential for productivity and for nutrient and 

water cycling partially determine the appropriate land-use and management practices 

for a site. Land-use practices that fall within these limits are usually cost-effective in 

terms of human resources and future costs caused by unwarranted changes on the 

land. Nevertheless, supplementing the natural resources of an area by adding 

nutrients through fertilization or water via irrigation is common. Even with such 

supplements, however, cost-effective management recognizes natural limitations of a 

site. Implementing land-use and -management practices that are compatible with the 

natural potential of the area requires that land managers understand a site’s potential. 

For example, land-management practices such as no-till farming reduce soil erosion 
142

or mitigate other resource losses. Often, however, land uses ignore site limitations or 

externalize site potential. For example, building shopping malls on prime agriculture 

land does not make the best use of the site potential. Nevertheless, land products are 

limited by the natural potential of the site. 

Together these guidelines form the basis of a landscape design perspective that 

should improve the ability to understand and manage the complex system that is affecting 

hypoxia in the Gulf of Mexico. 

Key Findings and Recommendations 

 

The large-scale policy models that have been developed to date each have 

strengths and weaknesses. None of the models adequately address the full range of 

management options (wetlands, buffers, nutrient management, etc.) or the full range of 

policy instruments in a geographically explicit manner. In fact, no single model is likely 

to be adequate for the full range of decision making that adaptive management of this 

complex system requires. Moreover, the focus of prior analyses was on cost-effective 

strategies to reduce N loss, which was the concern at the time. Given that the best current 

science suggests P is also a limiting nutrient in the Gulf, it is important to seek costeffective practices that affect both N and P while considering possible tradeoffs between 

them. 

 The CENR study remains the only research effort to consider the overall costs and 

benefits of controlling hypoxia in the Gulf of Mexico. The study suffers from a number 

of shortcomings (many control options and sources of nutrients were not considered, the 

hydrology of fate and transport was ignored, and no sensitivity analysis concerning key 

assumptions was undertaken to name a few). The evidence from this work and other 

studies suggests that it is probable that social welfare in the basin can be maintained 

while achieving the goal of a 5-year running average of 5000 km2

 for the hypoxic zone. 

Most importantly, welfare losses from costs incurred to control hypoxia in the Basin will 

be offset, at least in part, by co-benefits of nutrient reductions. For example, research on 

wetlands in the MARB suggests that the benefits of large scale restoration efforts would 

exceed the costs. Second, only limited targeting of control options that focus on hypoxia 

reduction and its co-benefits have been undertaken. Given the significant gains in cost 

savings that targeting can achieve, this suggests that it may be possible to achieve 

hypoxia reduction at lower cost than predicted in models that do not consider complete 

targeting. Based on these findings, the SAB Panel offers the following 

recommendations. 

• The management of factors affecting hypoxia within the MARB should be viewed 

as components of a designed landscape so that costs and benefits at various spatial 

and temporal scales are explicitly considered. 

• Integrated economic and watershed models are needed to support an adaptive 

management framework. Models are needed that represent land use and costs of 
143

conservation at both the fine scale, such as the 8 or 12-digit HUC size, as well as 

a larger scale that encompasses the entire MARB. 

• Research that assesses the optimal suites of conservation practices to maximize 

both local water quality and other co-benefits and Gulf hypoxia reduction is 

needed. This will require improved understanding of the watershed scale benefits 

of these control measures and their costs. 

• To reduce hypoxia and protect social welfare in the MARB, control measures that 

both reduce hypoxia cost-effectively and provide co-benefits in the MARB should 

be targeted whenever possible. Targeting control measures can reduce the costs 

and increase co-benefits associated with measures to control hypoxia in the Gulf 

of Mexico. 

4.4. Cost-Effective Approaches for Non-point Source Control 

 While the Action Plan and this Advisory urge the reliance on adaptive 

management principles, a variety of tools can be used as the vehicle for implementation 

within adaptive management. The current Action Plan indicates a principle of 

encouraging “actions that are voluntary, practical, and cost-effective” (page 9). 

Additionally, the plan will “utilize existing programs, including existing State and 

Federal regulatory mechanisms,” as well as identify needs for additional funding. These 

statements include a variety of tools ranging from purely voluntary programs (those with 

no associated financial incentives) to current conservation programs funded by state and 

federal agencies (such as the Conservation Reserve Program (CRP) and the 

Environmental Quality Incentive Program (EQIP)) to water quality trading. Research 

assessing the costs and effectiveness of these approaches is addressed in this section. 

 Complicating the design of cost-effective approaches is the geographic distance 

between the sources of nutrients and the receiving waters downstream. Two identical 

farm fields in different locations (with resulting differences in the hydrology of the local 

watershed) will send differing amounts of nutrients to the Gulf. Hence, the effectiveness 

of a practice or sink in a particular location depends on what sources and sinks are 

present elsewhere in the watershed. Whether it is cost effective to install a buffer at a 

particular location may depend upon whether there is a wetland at the base of the 

watershed, whether conservation tillage is being practiced elsewhere, etc. Thus, rather 

than focus on individual practices, policy options that can simultaneously encourage the 

adoption of practices and sinks that are jointly cost effective will best protect social 

welfare in the Basin. 

 It is important to clarify the concept of “costs.” Here, “costs” refers to the least 

amount of compensation needed to effect change, e.g., the compensation that would be 

necessary for a landowner or farmer to adopt a conservation practice. This is the standard 

concept of economic cost, relevant to any good or service. This cost includes “direct” 
144

costs such as the cost of new equipment, building of structures, and labor to manage a 

practice, as well as a myriad of potential “indirect” costs such as lost profits from 

adopting the practice, compensation for added risk from the practice, etc. Components of 

these costs can be negative; i.e., it may actually increase profitability to adopt some 

practices (conservation tillage in certain circumstances is a notable example). 

 Second, the focus of most economic studies is on total costs with little or no 

consideration paid to what subset of society actually bears the costs (incidence) of the 

policy. This focus on efficiency (seeking the lowest cost approach) is based on the 

premise that compensation could always be paid to those bearing the cost in some form 

so that society will be best off if the lowest cost option is pursued. However, since such 

compensations are rarely paid, the issue of who pays is likely to enter the policy decision. 

Complete information on the incidence of alternative tools in this context is not available, 

but where appropriate, we note the likely incidence considerations. 

4.4.1. Voluntary programs – without economic incentives 

 There is a small and growing literature concerning the effectiveness and optimal 

design of voluntary agreements that do not have positive or negative financial incentives 

associated with them (National Research Council, 2002; Morgenstern and Pizer, 2007). 

Key insights were presented in a game-theoretic model by Segerson and Miceli (1998), 

who identified the conditions under which voluntary agreements are likely to yield 

efficient pollution levels without significant economic incentives. They studied 

voluntary agreements that are based on threats of harsher outcomes if the goals are not 

met, using the example of mandatory abatement requirements if the voluntary agreement 

does not succeed in meeting the pollution goal. The premise is that firms will voluntarily 

agree to reduce pollution if they can avoid the costs that future mandatory controls would 

otherwise bring. In the absence of financial compensation, the presence of a positive 

probability of a penalty (or cost in the form of mandatory control) is required to support 

Segerson and Miceli’s findings that there are situations in which efficient levels of 

pollution control can be achieved with voluntary agreements (without economic 

incentives). They found that pollution reduction is likely to be small when the 

background threat is weak. 

 Empirical work also sheds light on the efficacy of voluntary agreements that do 

not have financial incentives. Mazurek (2002) identified 42 voluntary environmental 

initiatives sponsored by the federal government since 1988. Although the programs she 

identifiee are largely outside the realm of agriculture, her conclusions are relevant. 

Mazurek concluded that a variety of implementation problems have led to “lower-thanexpected” environmental results for voluntary (without financial incentive) agreements, a 

result consistent with findings of a 1997 USGAO (1997) report concerning four voluntary 

agreements related to climate change. 

 In the same National Research Council report (2002), Randall identified three 

essential functions for government if voluntary agreements (without financial incentives) 
145

are to be effective. These key functions are meaningful monitoring to back up a threat of 

government inspection, “credible threat of regulation” if the goals are not met, and a clear 

liability system to punish “blatant polluters and repeat offenders.” Randall concluded 

that “voluntary (or negotiated) agreements, industry codes, and green marketing should 

be viewed as promising additions to the environmental toolkit, but they should 

supplement, not supplant, the regulatory framework. They make a nice frosting on the 

regulatory cake. But the cake itself must be there (pages 317-318).” 

 Finally, Morgenstern and Pizer (2007) presented seven case studies on voluntary 

agreements (without economic incentives) in the U.S. and elsewhere. Point estimates of 

environmental improvements attributable to the voluntary programs ranged from negative 

values (actual declines in environmental performance) to a maximum of 28% 

improvement in environmental performance. Morgenstern and Pizer concluded “that 

voluntary programs have a real but limited quantitative effect... (page 182).”

 Given the historical aversion to imposing mandatory requirements in agriculture, 

the collective weight of these studies suggest that voluntary agreements that do not have 

incentives associated with them are not likely to be adequate on their own to achieve 

significant reductions in nutrient runoff. In short, voluntary programs without incentives 

can have small effects but cannot be relied upon to induce major environmental 

improvements. 

4.4.2. Existing Agricultural Conservation Programs 

 Currently, the largest incentive-based conservation programs related to agriculture 

are the EQIP and CRP. A potentially significant program introduced in the 2002 Farm 

Bill was the Conservation Security Program (CSP), which has been funded only partially 

and implemented incrementally. The CRP pays farmers to retire land, and the other two 

pay farmers to implement conservation practices on their farms (EQIP is a cost-share 

program; CSP was intended to cover the full costs of adoption). Numerous studies 

undertaken by USDA’s Economic Research Service and others have estimated the 

magnitude of environmental benefits from these programs in physical terms (e.g., tons of 

erosion reduction, acres of habitat preserved, acres of wetlands restored, etc) and some 

efforts have been made to monetize these benefits [see Claassen et al. (2004) for a 

summary of CRP studies as well as Haufler (2005)]. The Conservation Effects 

Assessment Program (CEAP) was initiated in an attempt to provide nationwide estimates 

of the benefits provided by the full suite of conservation programs; a national assessment 

of the water quality benefits is being developed currently (Bob Kellogg, presentation to 

SAB Hypoxia Advisory Panel, December 6, 2006). 

 The CRP pays landowners to take their land out of crop production and place it in 

perennial vegetation or trees, depending on the region of the country, with a goal of 

creating wildlife habitat and reducing erosion (and originally to reduce crop production). 

The CRP enrolls about 10% of total US cropland, nearly all in ten-year contracts 

although there is significant concern that high corn prices due to ethanol expansion may 
146

rapidly reduce this amount. A number of studies have identified large environmental 

benefits associated with the CRP [Smith and Alexander (2000), Feather et al. (1999)]. 

The program has used an Environmental Benefits Index (EBI) since 1990 to prioritize 

parcels for inclusion in the program that gives points to land based on particular 

environmental attributes and cost. The movement from targeting erodible lands (prior to 

1990) to the use of the EBI for targeting has been estimated to have doubled the benefits 

from the program (Feather et al., 1999). Ribaudo (1989) estimated that a CRP enrollment 

that targets lands based on environmental damages (benefits) would have significantly 

greater benefits still. By redesigning the weights in this index, the program could target 

land that is predicted to contribute high nutrient loadings to the Gulf. 

 Many other studies have addressed the cost-effectiveness of land retirement to 

achieve environmental benefits within the context of the CRP. In a series of papers 

assessing the efficiency of the Conservation Reserve Enhancement Program (CREP) in 

Illinois, Khanna et al. (2003) linked the AGNPS model with site specific characteristics 

of parcels to examine the relative efficiency of alternative targeting mechanisms (Yang et 

al., 2003, 2004, and 2005). Extremely large gains from targeting were reported; for 

example, Yang et al. (2004) estimated that with targeting, 30% less cropland could have 

been retired (at almost 40% less total cost) while achieving 20% reductions in erosion 

instead of the actual 12% reduction. 

 The EQIP program is a cost share program for conservation practices in livestock 

facilities and on land that remains in agricultural production. A prospective benefit cost 

analysis (as required by Executive Order 12866) predicted over $5 billion in net benefits 

from the EQIP program as implemented under the 2002 Farm Bill, even though not all of 

the benefits could be monetized (US Department of Agriculture, 2003). 

 The Wetland Reserve Program (WRP), Grassland Reserve Program (GRP), and 

Wildlife Habitat Incentive Program (WHIP) are all smaller land retirement programs that 

also could potentially benefit efforts to reduce Gulf hypoxia. Additional information on 

the large-scale potential for wetlands is provided by Hey et al. (2004), who addressed the 

question of whether the social benefits from restoring up to 2.83 million hectares (7 

million acres) of cropland in the 100 year floodplain of the upper Mississippi River basin 

to wetlands exceed the costs. The benefits include reduced flood related crop damages, 

reduced crop subsidies and non-flood related recreation benefits of wetland conversion 

including fishing, hunting, and general recreation usage. These benefits were compared 

to estimates of the costs of cropland conversion comprised of farm rental rates 

(representing the present value of farmland income) and the costs of wetland construction 

and maintenance. Hey et al. (2004) estimated that the benefits exceed the costs in all 

locations considered except one county in Missouri. In the context of NGOM hypoxia, 

this difference is especially striking because the benefits exceed the costs for this 

conversion even without considering any benefits from reduction of the hypoxic zone. 

As the authors carefully pointed out, the social efficiency of converting 2.83 million 

hectares (7 million acres) does not mean that private benefits will exceed the private costs 

for all parties. Individual landowners would stand to lose while recreationists accrue 

benefits. 
147

 These findings represent an important addition to the assessment of wetlands in 

the Integrated Assessment. While Doering et al. (1999) concluded that wetland 

restoration was less cost-effective than fertilizer reductions, their analysis did not include 

cost savings from crop subsidy reductions nor flood related crop damages. In addition, 

the Hey et al. (2004) work focuseD on wetlands targeted in flood plains. The study 

suggests two points of key importance for NGOM hypoxia; 1) there is a large amount of 

acreage that is situated in locations that potentially could serve as nutrient sinks in the 

upper Mississippi River basin, and 2) the co-benefits of this action are large enough, in 

and of themselves, to justify the social efficiency of converting this land to nutrient sinks 

even without considering the benefits associated with reducing Gulf hypoxia. 

 The programs mentioned above can be categorized into one of two groups: land 

retirement programs and “working” land programs. Both the CRP and WRP are 

examples of land retirement programs, since landowners receive payments in exchange 

for taking land out of active agricultural production and putting the land into perennial 

grasses, trees, or wetlands restoration. In contrast, EQIP and the CSP are examples of 

working land programs whereby landowners or producers receive payments to cover part 

or all of the costs of making changes in conservation practices or management decisions 

on their land that remains in agricultural production. Some research has addressed the 

cost-effectiveness of working land programs vs. land retirement programs. For example, 

Feng et al. (2006) found that a cost-effective allocation of resources to sequester carbon 

in agricultural soils favors working land (via conservation tillage subsidies) over land 

retirement (via payments to retire land and plant it in perennial grasses). It is important 

to note however, that this study focused on stylized working land and land retirement 

programs rather than attempting to address the cost-effectiveness of existing conservation 

programs as actually implemented. 

 The existing working land and land retirement programs are implemented with 

features that likely affect the cost-effectiveness of the programs for achieving 

environmental gains in different ways. For example, the CRP uses an EBI that favors 

admitting land into the program that achieves environmental benefits at relatively low 

costs. All else equal, this component of the program will improve its cost-effectiveness. 

In contrast, the CSP provides payments for ongoing stewardship of farmers so that 

program expenditures are used to reward past behavior rather than to change existing 

behavior. This, all else equal, will reduce the program cost-effectiveness for achieving 

environmental gains. The lack of competitive bidding and clear targeting also reduces 

the cost-effectiveness of this program. Finally, it is worth noting that targeting and 

competitive bidding were explicitly disallowed in the EQIP program during its last 

reauthorization. Again, this will reduce its cost-effectiveness. 

4.4.3. Emissions and Water Quality Trading Programs 

 Emission trading is a regulatory approach that sets a maximum allowable level of 

overall emissions and then allows sources to exchange pollution allowances. A properly 
148

structured trading program can reduce the costs of achieving emission standards by 

allowing the flexibility necessary to focus pollution reductions on sources that are less 

expensive to control. In theory, a broad based emissions trading program could help to 

reduce the air and water contributions of nutrients to the NGOM. Water quality trading is 

simply the name given to the extension of emissions trading to achieving water quality 

objectives. 

 In a recent survey of the programs to support water quality trading in the U.S., 

Breetz et al. (2004) identified 40 water trading initiatives and an additional six state 

policies with specific programs related to water quality trading. EPA has supported these 

programs (US EPA, 2004a) and has produced explicit policies related to their 

implementation. Many states and regions also have explicit policy guidance. However, 

the effectiveness of these programs appears to have been quite limited as very few trades 

are actually occurring. Further, little evidence of environmental improvement associated 

with these programs exists (Breetz et al., 2004). 

 A key problem with these programs is the lack of a required water quality 

improvement necessary to generate adequate demand for credits (King, 2005). To 

achieve “cap and trade,” an effective cap is necessary. A cap could come from a tight 

enough cap on point sources such that they would find it cost-effective to purchase 

credits from agricultural non-point sources. Alternatively, the cap could be extended to 

agricultural sources. While some have conjectured that the Total Maximum Daily Load 

(TMDL) program may eventually play this role, there is no current mandate for 

agricultural sources to restrict nutrient runoff. Also problematic are a range of 

restrictions on allowable trading such as requirements that a particular baseline set of 

conservation practices be in place with credits accruing only for additional conservation 

activity. 

 While trading could be a significant contributor to cost-effective nutrient control, 

the necessary institutions for water and/or air emissions trading to be an effective policy 

instrument are not broadly in place. In addition to clear and enforceable limits on 

emissions or water quality contributions (from point and/or non-point sources), 

enforceable rules concerning trading ratios, liability when standards are not met, 

monitoring, etc. must be established before these markets can flourish. Ideally, a trading 

program to address NGOM hypoxia would be broad based and include highly diverse 

sources (such as air deposition and many agricultural non-point sources) to maximize the 

potential for cost savings. 

4.4.4. Agricultural Subsidies and Conservation Compliance Provisions 

 U.S. farmers have been the recipients of farm payments for decades. These 

payments support prices and/or income, especially of farmers growing bulk commodities 

such as corn and soybeans. Economic theory suggests that, all else equal, such payments 

will increase the intensity and acreage of farming, possibly resulting in increased water 

quality problems. Research by Reicheldorfer (1985) provided empirical evidence that 
149

these payments encourage crop production on highly erosive land. Likewise, a recent 

study from USDA’s Economic Research Service (Lubowski et al., 2006) quantified the 

effect of one major program, subsidized crop insurance, on the location and acreage of 

cropland and its environmental effects. Lubowski et al. (2006) estimated that about a 

million hectares (2.5 million acres) were brought into production as a result of the 

program and that these lands are more vulnerable to erosion, are more likely to include 

wetlands, and have higher levels of nutrient losses than average. 

 To some extent, USDA’s conservation programs (see Section 4.4.2) exist to 

counteract the “perverse effects” or unintended consequences of its crop subsidies 

inasmuch as government financial support has encouraged farmers to choose commodity 

crops that require more fertilizer, maximize yield without regard to soil and water quality 

consequences, and cultivate marginal land. Re-structuring or eliminating existing 

subsidies could serve to mitigate some of these perverse effects (e.g., by shifting 

subsidies to reward less fertilizer-intensive crops as well as by requiring, as a condition of 

receiving subsidies, certain conservation practices). 

 Taheripour et al. (2005) provided additional evidence on this point. First, their 

model suggests that removal of all crop subsidies would reduce nitrogen pollution by 

8.5% and that the reduced need for distortionary income taxes to support these subsidies 

could increase social welfare by $1.2 billion. Further, they found that tax neutral policies 

to achieve nitrogen reduction can generate significant double dividends (a double 

dividend refers to a situation where a policy not only internalizes an externality but also 

reduces the deadweight losses associated with distortionary taxation such as an income 

tax). They provide an estimate of the magnitude of the double dividend for a range of 

nitrogen reduction goals and policy approaches including a nitrogen tax, a nitrogen 

reduction subsidy, a tax on output, and a combined output tax and nitrogen reduction 

subsidy and find that a double dividend from these instruments can be significant. 

 While environmental improvements associated with agriculture have largely been 

pursued via cost-share or subsidy programs, one significant regulatory approach has been 

the implementation of environmental compliance provisions that require farmers who 

receive farm program payments (including price support and income support) to 

undertake some environmental performance practices. Specifically, in the 1985 Food 

Security Act, conservation compliance provisions required owners of highly erodible 

land (a categorization of land based on its slope and soil type) to implement soil 

conservation plans and a “swampbuster” provision disallowed payments to go to farmers 

who converted wetlands to crop land. Claassen et al. (2004) estimated that up to 25% of 

the reduction in soil erosion that occurred between 1982 and 1997 was attributable to 

conservation compliance. Many believe these gains could have been higher if there had 

been stronger enforcement of the mechanism. While no direct estimates are available of 

the increased benefits that could come from more enforcement, there is evidence of very 

limited reporting and penalizing of violations (Claassen, 2000). 

 Claassen et al. (2004) assessed the prospect for reducing nutrient losses from the 

Mississippi River basin by extending compliance requirements to nutrient management. 
150

They used “nutrient management” to refer to the range of activities related to the timing 

and level of fertilization decisions that best minimizes soil nutrients in excess of crop 

needs at any point in time. They noted that the ideal set of nutrient management practices 

will vary considerably across farms and regions and that the costs of these activities will 

also vary notably across this space. Using data from the EQIP program, they summarized 

the distribution of incentive payments needed to induce willing adoption of nutrient 

management practices as defined under EQIP. For the Heartland region (ERS Farm 

Resource Region), the average annual incentive payment is about $7 per acre, and 95% of 

the payments are $12 per acre or less. 

 While these data provide an excellent starting point for assessing the cost 

effectiveness of nutrient management methods addressing local water quality and NGOM 

hypoxia, several additional pieces of information would be needed for a full assessment. 

First, these costs represent the compensation needed for those farmers who have already 

adopted practices under the EQIP program; those who have not adopted are likely to have 

at least as high costs, possibly substantially higher. In this regard, these costs could be 

viewed as a lower bound. Second, these costs are specific to the EQIP requirements for 

nutrient management. Whether these requirements are effective enough to yield 

substantial off-site benefits is not addressed. Nonetheless, based on this cost assessment 

and a comparison with the annual commodity program payments farmers typically 

receive, Claassen et al. (2004) concluded that substantial nutrient management could 

occur with extension of conservation compliance provisions to nutrients. 

 Claassen et al. (2004) also considered whether buffer practices could be induced 

under conservation compliance provisions. They included riparian buffers, filter strips, 

grassed waterways, and contour grass strips in their discussion of buffer practices. To 

assess the costs of these practices and how they vary across locations, they looked at 

information on producers’ willingness to accept compensation for adoption of the 

practices from the priority areas sign up of the continuous CRP. Owners of these lands 

received an average payment of about $90 per year in addition to 50% cost share for 

installation of the buffer practice. Based on this analysis, as an example, Claassen et al. 

(2004) computed the annual costs per area for a filter strip and concluded that, in many 

cases, this payment would be below the average subsidy received by producers, thereby 

suggesting that buffer practices might also be successfully adopted under nutrient 

compliance provisions. 

 Finally, Claassen et al. (2004) noted that conservation compliance provisions are 

likely to have few transaction costs relative to other policies (although enforcement costs 

would need to be considered) and require very low budgetary outlays beyond the 

payments that are already provided for commodity or insurance programs. Claassen et al. 

(2004) also argued that conservation compliance requirements have been relatively costeffective due to the flexibility with which they can be implemented. Producers in 

different regions of the country, with differing soil and weather conditions, can meet their 

compliance obligations with different practices. This flexibility means that the most 

appropriate technologies can be used for the location of the practice. 
151

4.4.5. Taxes 

 The use of a per unit tax to internalize the costs of externalities of production is 

well known to be highly cost effective when the tax is placed directly on the externality 

generating activity; these “Pigouvian” taxes are the equivalent of placing the appropriate 

price on the pollutant (Baumol and Oates, 1988). Taxes can be a powerful market signal, 

communicating the need to change behavior, Baumol and Oates (1988) demonstrated that 

subsidies (essentially just negative taxes) can also be designed that provide the equivalent 

market signals for changes in behavior. This argument is often used to support the design 

of environmental programs that pay participants for the provision of environmentally 

friendly practices rather than using taxes to change behavior. A potentially important 

exception to this equivalence can occur when the provision of a positive payment induces 

entry into the farming sector generating production on otherwise unprofitable lands. This 

possibility was addressed in Section 4.4.4 in the context of general agricultural subsidies 

and conservation compliance. 

 A tax directly on an input into production that is highly correlated with the 

pollutant can be an efficient second-best policy. The possible use of a nitrogen fertilizer 

tax was considered in Doering et al. (1999) and found to be as cost-effective as any of the 

policies they considered (they note that the initial incidence falls on farmers). Fertilizer 

taxes already exist in some states, but are set at much smaller levels than those studied by 

Doering et al. (1999). The inelastic demand for fertilizer (Denbaly and Vrooman, 1993) 

means that the magnitude of taxes needed to induce behavioral change would likely be 

large. 

 The incidence of a tax (and thus determination of who pays the costs) is likely to 

fall on farmers and consumers of food products made from crops that use fertilizer. In 

contrast, the incidence of conservation program payments is largely on taxpayers. 

Finally, it is important to note that tax instruments will be more efficient the more 

broadly they are applied to the various nutrient sources identified as pollutant 

contributors; so ideally a tax would be applied to all nutrient sources rather than singly to 

fertilizer. 

4.4.6. Eco-labeling and Consumer Driven Demand 

 The idea that environmentally friendly producer behavior can be induced by 

consumer demand is one basis for eco-labeling and certification programs. Dolphin safe 

tuna (Teisl et al., 2002) and organic fruits and vegetables (Loureiro et al., 2001) are two 

successful examples. Research analyzing the effectiveness of eco-labeling suggests some 

promise. 

 Thogersen (2002) summarized three schemes, all implemented in Europe, that 

have been credited with significant reductions in emissions from heating appliances and 

paint solvents (the German “Blue Angel” brand) and reductions in pollutants from paper 
152

production and household chemical and laundry emissions (the Swedish “Good 

Environmental Choice” label and the Nordic “Swan” label). Although not specific to a 

particular product, Clark and Russell (2005) noted that several studies of the Toxic 

Release Inventory have shown that information can affect firms’ choices. 

 Could consumer driven demand affect the changes in land use and agricultural 

management necessary to contribute notably to nutrient flows into the Gulf? This 

approach would require the labeling of food and fiber products made from agricultural 

outputs in the MARB to indicate that they were produced in such a way as to reduce or 

eliminate nutrient contributions to hypoxia. Consumers would then need to respond to 

this labeling by purchasing products, presumably at a higher cost, in adequate quantity to 

change the market behavior. Given that much of the grain produced in the Corn Belt is 

used for livestock feed and not directly traceable to its field of origin, it will be difficult 

to distinguish products that were produced with “hypoxia-friendly” production practices 

from those that were not. It is not clear that labeling can credibly be produced without 

significant government involvement and expense (Crespi and Marette, 2005). Nor is it 

clear that consumer response would be adequate to drive changes in production practices, 

even if the labeling challenges could be overcome. One area in which labeling may 

prove effective is in animal agriculture where the tracking of an individual unit from 

producer to final consumer is more straightforward. 

4.4.7. Key Findings and Recommendations on Cost Effective Approaches 

 Voluntary agreements with no accompanying economic incentives are not likely 

to be adequate to obtain significant reductions in N and P. While there may still be some 

low-cost conservation practices that can be implemented in some locations (better 

“crediting” for manure spreading for example), nutrient reductions that face agricultural 

producers with costly tradeoffs cannot be expected without strong economic signals. 

These economic incentives can take many forms: conservation payments such as those in 

many current agricultural conservation programs, taxes, restructuring or removal of 

subsidies (such as conservation compliance provisions), etc. 

Water quality trading programs have not yet demonstrated the ability to improve 

environmental performance and/or reduce costs of meeting environmental targets 

primarily due to an absence of effective emissions restrictions. However, with clearer 

water quality improvement mandates and more flexible rules for trading, these programs 

could develop into cost-effective instruments. 

Numerous studies have demonstrated that existing incentive-based conservation 

programs, specifically the CRP, WRP and EQIP, have provided significant 

environmental benefits. However, these programs can be much more cost-effective with 

additional targeting and competitive bidding mechanisms. Given the menu of existing 

programs, it is possible to reduce hypoxia and protect water quality in the MARB without 

significant new government funding, although the distributional consequences of the 

various approaches will differ. Based on these findings, the SAB Panel offers the 

following recommendations. 
153

• To achieve N and P reductions from agricultural sources of the magnitude needed 

to affect hypoxia, economic incentives are needed to induce adequate adoption of 

conservation practices. These incentives can take many forms: conservation 

payments, taxes, and/or restructuring of existing farm subsidy and compliance 

requirements. 

• To maximize the N and P reductions achieved with federal and state conservation 

dollars (e.g., CRP, WRP and EQIP), targeting and competitive bidding 

mechanisms are needed so that lands enrolled in these programs achieve 

maximum environmental benefits at lowest cost. Strategically placed wetlands in 

the upper Mississippi River basin could serve as effective nutrient sinks. 

Research has demonstrated that the local co-benefits are large enough, in and of 

themselves, to justify restoring these wetlands. The additional benefits associated 

with reduction in Gulf hypoxia reinforce the conclusion of the desirability of 

wetlands restoration. 

• Water quality trading programs hold promise, but, without enforceable caps 

(water quality standards), these programs cannot be expected to achieve much 

nutrient reduction. 

• To minimize the adverse effects of existing agricultural subsidy programs, 

conservation compliance requirements that target reductions in nutrients could be 

very cost-effective, but only with adequate enforcement. 

• To select policies and programs with maximum economic efficiency, all cobenefits should be considered regardless of which policy tools are used. For 

example, since wetlands provide valuable habitat and flood control in addition to 

water quality benefits, there may be instances in which it is desirable to control 

nutrients by restoring wetlands, even if it is less costly to reduce nutrients by 

managing croplands. 
154

4.5. Options for Managing Nutrients, Co-benefits, and Consequences 

4.5.1. Agricultural drainage 

 The Integrated Assessment reports identified several research needs related to 

agricultural drainage. Brezonik et al. (1999) emphasized the importance of agricultural 

drainage in nutrient transport from cropland and identified increased spacing of 

subsurface drainage tile and controlling water table levels (controlled drainage) among 

those practices that could potentially reduce nitrate losses from cropland. Mitsch et al. 

(1999) noted that controlled drainage was not widely practiced in US Corn Belt and that 

most of the research on controlled drainage had been conducted in more southern 

climates. 

Alternative drainage system design and management 

 Relatively few field studies have addressed the effects of subsurface drain depth 

and spacing on N losses from cropland. Overall, results suggest a trend of decreased 

subsurface flow and decreased N loss at wider tile spacing or decreased tile depth. 

Reported reductions in nitrate export are primarily due to reductions in the volume of 

flow rather than reductions in nitrate concentration. Drain flows and N loss can be 

affected by both drain spacing and depth (Hoffman et al., 2004; Kladivko et al., 2004; 

Skaggs and Chescheir, 2003; Skaggs et al., 2005), and use of drainage intensity (Skaggs 

et al. 2005) normalizes some of the variability in results of drainage spacing studies. 

Drainage intensity increases with deeper tile depths and closer tile spacing. Research 

suggests that reducing drainage intensity by either shallower tile depth or wider tile 

spacing will reduce subsurface flow and nitrate loss. However, adjustments in tile 

spacing and depth are only possible when drainage systems are being installed, and the 

Corn Belt is already extensively drained. As these systems are replaced, repaired, and 

upgraded over the next few decades, there will be opportunities to consider alternative 

drainage designs to minimize nutrient losses. In the meantime, there may be 

opportunities to achieve similar benefits by retrofitting existing drainage systems with 

control structures that allow some management of subsurface drainage. 

 Drainage management (controlled drainage) is currently an area of active research 

and development (http://extension.osu.edu/~usdasdru/ADMS/ADMSindex.htm). 

Research suggests that drainage management could reduce nitrate transport from drained 

fields by 30% for regions where appreciable drainage occurs in the fall and winter 

(Cooke et. al., in press). Although water table management could potentially alter 

nitrification and denitrification reactions, reported reductions in nitrate export with 

controlled drainage are primarily due to reductions in the volume of flow rather than 

reductions in nitrate concentration. Some uncertainty arises from difficulties in closing 

water balances (and therefore N balances) in field studies, and an unknown amount of 

subsurface flow reduction could be due to lateral seepage and/or increased surface runoff 

(Cooke et al., in press). Simulation studies predict increased surface runoff when higher 

water tables are maintained using controlled drainage (Skaggs et al., 1995; Singh and 
155

Helmers, 2006) suggesting a potential tradeoff between reduced subsurface drainage and 

increased surface runoff. Although raising the water table can decrease the volume of 

infiltrating water entering drainage tile, higher water tables can also increase surface 

runoff resulting in increased erosion and loss of particulate contaminants such as soil 

bound phosphorous. 

 Controlled drainage requires relatively flat and uniform topography, and slopes of 

less than 0.5% or 1 % are recommended (Cooke et al., in press; Frankenberger et al., 

2006). Concerns for erosion and surface runoff increase with increasing slope, and 

slopes greater than 0.5-1% can require an impractical number of control structures. 

There has been speculation that new technologies could make the practice economically 

feasible at slopes of 2% or more, but this would raise even greater concerns over surface 

runoff. Although tile drainage is widespread throughout the Corn Belt, it is not clear 

what portion of this tile drainage can be retrofitted with structures for controlled drainage. 

A first approximation might be an estimate of the fraction of tile drained lands with 

slopes less than 0.5-1%, but this approach requires higher resolution topography than is 

generally available in the Corn Belt. These estimates are available for a few large 

drainage districts in north central Iowa for which very high resolution topography were 

developed. Although 50 to 75% of the cropland in these drainage districts is tile drained, 

only about 10% has a slope less than 1% and only about 3% has a slope less than 0.5% 

(Matt Helmers, Iowa State University, Ag Drainage Website, 

http://www3.abe.iastate.edu/agdrainage). These results suggest that controlled drainage 

may be applicable to a relatively small fraction of tile drained land in Iowa, but this may 

not be representative of other regions of the Corn Belt. Based on STATSGO soils data, 

Illinois, Indiana, and Ohio may have twice as much cropland suitable for controlled 

drainage as Iowa (Dan Jaynes, National Soil Tilth Lab, Ames, IA). High resolution 

topography could provide a much better basis for this assessment. 

Bioreactors

 Denitrification bioreactors have been installed in the field as treatment systems 

for tile drain effluent (Van Driel et al., 2006) and as denitrification walls (a trench filled 

with carbonaceous material to intercept subsurface flow) (Schipper and VojvodicVukovic, 1998; Robertson et al., 2000; Schipper and Vojvodic-Vukovic, 2001; Schipper 

et al., 2004; Schipper et al., 2005). Bioreactors on tile drains are typically bypassed 

during high flows and "are most usefully applied in the treatment of baseflows rather than 

peak flows." Current knowledge indicates that denitrification walls are effective for at 

least 5 to 7 years with little or no loss of nitrate removal capacity (Robertson et al., 2000; 

Schipper and Vojvodic-Vukovic, 2001). A variety of materials such as corn stalks, wood 

chips, and sawdust are potential organic amendments to enhance denitrification in 

bioreactors. Continued research is needed to determine whether denitrification 

bioreactors could be installed around lateral tile drain lines and whether this would be 

technically and economically feasible. Future re-design of tile drain systems may include 

integrated denitrification enhancements around tile lines and at the outlets of smaller tile 

lines. 
156

Key Findings and Recommendations: 

 Alternative drainage designs with reduced drainage intensity due to shallower tile 

depths and/or wider tile spacing could significantly reduce nitrate losses but can be 

expected to increase surface runoff and losses of particulate contaminants. Controlled 

drainage could significantly reduce nitrate losses where appreciable drainage occurs in 

the fall and winter but can be expected to increase surface runoff and losses of particulate 

contaminants. Controlled drainage is most appropriate for areas having slopes of less 

than 0.5-1%, and it is not clear what fraction of tile drained lands are suitable for 

application of controlled drainage. In some areas, slope could seriously constrain 

applicability of the practice. Bioreactors can significantly reduce nitrate concentrations 

but typically must bypass peak flows during which much of the nitrate load is 

transported. Based on these findings, the SAB Panel offers these recommendations. 

• Additional research is needed to evaluate topographic constraints on the 

applicability of controlled drainage including developing high resolution 

topography for the Corn Belt. 

• Additional research is needed to fully characterize water and nutrient balances for 

alternative drainage design and management most critically using small watershed 

scale studies (less than 2,500 hectares or about 10,000 acres) to document effects 

when scaled up. 

• A strategy for implementation of alternative drainage design or management 

should be developed that includes consideration of potential trade-offs between 

reduced nitrate loss through tile drains and increased P loss through surface 

runoff. 

4.5.2. Freshwater Wetlands 

 If wetlands are to serve as long-term “sinks” for nutrients, reductions in nutrient 

loads must reflect net storage in the system through accumulation and burial in sediments 

or net loss from the system, for example through denitrification or vegetation removal. 

The effectiveness of wetlands in reducing N export from agricultural fields will depend 

on the magnitude and timing of NO3 loads and the capacity of the wetlands to remove 

NO3 by denitrification. In contrast to NO3, gaseous losses of P are insignificant, and 

sediment accretion of bound inorganic P and unmineralized organic P is the primary 

mechanism by which wetlands serve as long-term P sinks. With the exception of P 

associated with suspended solids, wetlands are generally less effective at retaining P than 

at removing NO3 (Reddy et al., 1999). 
157

Nitrogen 

 The effectiveness of wetlands in NO3 reduction is a function of hydraulic loading 

rate, hydraulic efficiency, NO3 concentration, temperature, and wetland condition. Of 

these, hydraulic loading rate and NO3 concentration are especially important for wetlands 

intercepting non-point source loads. Hydrologic and NO3 loading patterns vary 

considerably for different landscape positions and different geographic regions. The 

combined effect of variation in land use, precipitation, and runoff means that loading 

rates to wetlands receiving non-point source loads can be expected to vary by more than 

an order of magnitude and will, to a large extent, determine NO3 loss rates for individual 

wetlands. 

 Mitsch et al. (2005a) examined NO3 retention in Mississippi River basin wetlands 

receiving non-point source NO3 loads either directly or through diversion of river water. 

Their study extended the earlier analysis of Mitsch et al. (1999) to include additional 

wetlands and to include wetlands outside the agricultural regions of the Corn Belt. They 

found that 51% of the NO3 mass reduction by the wetlands examined could be explained 

by a nonlinear regression based on annual mass load of NO3 per area of wetland. 

However, when the analysis is restricted to Corn Belt wetlands that receive seasonally 

variable water and nutrient loads (i.e., subjected to non-point source loading regimes), the 

relationship is much weaker (Crumpton et al. 2006, in press). Based on 34 “wetland 

years” of available data (12 wetlands with 1-9 years of data each) for sites in Ohio 

(Mitsch et al., 2005a; Zhang and Mitsch, 2000, 2001, 2002, and 2004), Illinois (Hey et 

al., 1994; Kovacic et al., 2000; Phipps, 1997; Phipps and Crumpton, 1994), and Iowa 

(Crumpton et al., 2006; Davis et al., 1981), percent mass NO3 removal is much more 

closely related to hydraulic loading rate (HLR) (Figure 41, R2

 = 0.69) than to mass 

loading rate (R2

 = 0.22). 

0 10 20 30 40

Hydraulic loading rate (m yr-1)

0

10

20

30

40

50

60

70

80

90

100

Percent mass nitrate remove

d
158

Figure 41: Percent mass nitrate removal in wetlands as a function of hydraulic loading rate. Best fit for 

percent mass loss = 103*(hydraulic loading rate)-0.33. R2 = 0.69. Adapted from Crumpton et al. (2006, in 

press). 

 Hydraulic loading rate explains relatively little of the variability in NO3 mass 

removal, which can vary considerably more than percent NO3 removal among wetlands 

receiving similar hydraulic loading rates. However, much of the variability in mass NO3

removal can be accounted for by explicitly considering the effect of HLR and flow 

weighted average (FWA) NO3 concentration (Crumpton et al., 2006, in press). For the 

wetlands in Figure 41, mass NO3 removal rate can be predicted as the product of percent 

removal (estimated as 103 * HLR-0.33) and mass load (estimated as HLR*FWA). This 

simplifies to the function [mass removal in kg N/ha/yr = 10.3 * (HLR in m/yr)0.67 * FWA 

NO3 concentration in g N/m3

] and explains 94% of the variability in mass NO3 removal 

for the wetlands considered here (Figure 42). The isopleths on the function surface in 

Figure 42 represent the combinations of HLR and FWA that can be expected to achieve a 

particular mass loss rate and illustrate the benefit of targeting wetland restorations in 

areas with higher NO3 concentrations. The wetlands examined by Mitsch et al. (2005a) 

had a median loading rate of 600 kg NO3-N/ha/yr, at which they predicted losses of 290 

kg NO3-N/ha/yr. This mass loss rate is near the lower mass loss isopleth of Figure 42 as 

would be expected for either low FWA concentrations at moderate to high HLRs or 

higher FWA concentrations at lower HLRs. Half of the wetlands considered by Mitsch et 

al. (2005a) had NO3 concentrations below 3 mg N/l. NO3 concentrations in tile drainage 

water commonly exceed 10 to 20 mg N/l (Baker et al., 1997, 2004, in press; David et al., 

1997; Sawyer and Randall, in press). The greatest benefit of wetlands for mass NO3

reduction will be found in those extensively row-cropped and tile-drained areas of the 

Corn Belt where NO3 concentrations and loading rates are highest. For these areas, NO3

mass removal rates could be several times higher than predicted by Mitsch et al. (1999, 

2005a). 
159

Figure 42: Observed NO3 mass removal (blue points) versus predicted NO3 mass removal (blue surface) 

based on the function [mass NO3 removed = 10.3*(HLR) 0.67 * FWA] for which R2

 = 0.94. Blue lines are 

isopleths of predicted mass removal at intervals of 250 kg ha/yr. The dashed, red line represents the 

isopleth for mass removal rate of 290 kg ha/yr suggested by Mitsch et al. (2005a). The green plane 

intersecting function surface represents organic N export. Adapted from Crumpton et al. (2006, in press). 

 Total and organic N data were available for about half of the wetlands represented 

in Figure 42. All of these wetlands were sinks for total N, but most were net producers of 

organic N, although in comparatively small amounts (and none were net producers of 

NH4). On average, FWA organic N discharged from the wetlands increased by 

approximately 0.2 g N/m3

 (range from <0 to 0.3 g N/m3) relative to incoming 

concentrations, with no relation to HLR or NO3 concentrations. The mass export of 

organic N was small compared to NO3 removal and had relatively little impact on 

reductions in total N, especially at higher NO3 concentrations. For comparison to mass 

NO3 loss, mass organic N export can be estimated as the product of HLR and the increase 

in FWA organic N and is represented by the green plane intersecting the function surface 

in Figure 42. At elevated NO3 concentrations, wetlands are nearly as effective in 

reduction of total N as in reduction of NO3. At very low NO3 concentrations, organic N 

production could equal NO3 removal, in which case wetlands would not function as total 

N sinks. 

 There is some concern over increased N2O emissions in wetlands exposed to high 

nitrate loads, and N2O emissions do increase in wetlands at elevated nitrate levels. 

However, N2O accounts for a very small fraction of N removal in wetlands receiving non 

point source nitrate loads, and N2O emission rates from these systems are very low 

(Hernandez and Mitsch 2006; Paludan and Blicher-Mathiesen 1996; Stadmark and 

Leonardson 2005). N2O emission accounted for only 0.3% of total N loss in wetlands 

receiving river flows with elevated nitrate levels (Hernadez and Mitsch 2006), and less 
160

than 0.13% of total nitrate loss in a wetland recharged by GW with elevated nitrate levels 

(based on maximum flux rates reported by Paludan and Blicher-Mathiesen 1996). N2O 

emission rates in wetlands receiving non point source nitrate loads average around 1 

umole N2O m-2 hour-1 (Hernandez and Mitsch 2006; Paludan and Blicher-Mathiesen 

1996) which is very similar to rates reported for cultivated crops in the Midwest (1-2 

umole N2O m-2 hour-1 (Parkin and Kaspar 2006; Grandy et al. 2006). The available 

research suggests that wetlands restored on formerly cultivated cropland for the purpose 

of nitrate removal would have little or no net effect on N2O emissions. 

Phosphorus 

 P removal in wetlands is controlled by three sets of processes: 1) sorption or 

release of P by existing sediments, 2) accumulation of P in new biomass, and 3) 

accumulation of P associated with the formation and accretion of new sediments/soils 

(Reddy et al., 2005). Existing sediments will have a finite capacity for sorption of P, 

determined in part by Al and Fe content in acid soils and by Ca and Mg content in 

alkaline soils. There will also be a finite capacity for the accumulation of P in new 

biomass. Of the three sets of processes, only the last contributes to long-term, sustainable 

P retention by wetlands: the accumulation of bound inorganic P and unmineralized 

organic P associated with the formation and accretion of new sediments and soil. 

 P sorption on both antecedent and newly accreting wetland soil is largely 

controlled by Fe, Al, and Ca. Reducing conditions found in wetlands may decrease 

sorption of P as insoluble complexes formed with Fe+3 are released upon reduction to 

Fe+2, solubilizing the P (Patrick et al., 1973). High S levels may enhance P flux from 

soils due to the binding of iron by sulfides (Bridgham et al., 2001; Caraco et al., 1989). 

Alkaline wetland soils are more conducive to P sorption than acidic wetland soils due to 

the presence of Ca in the alkaline wetland soils and the formation of insoluble Ca-bound 

P (Bruland and Richardson, 2006; Richardson, 1999). These two studies indicate that 

wetlands developed on soils rich in calcite and exchangeable Ca are likely to be more 

effective sinks for P under the reducing conditions necessary for denitrification. More 

research is needed to understand 1) the effects of wetland creation such as is being done 

in the upper Mississippi River basin and 2) whether wetlands created/restored on 

Mollisols will be effective P sinks due to formation on Ca-P complexes in addition to 

sedimentation and SOM formation. Bruland and Richardson (2006) determined that 

marshes with a higher soil P sorption index (amount of P sorbed by soil from a phosphate 

solution in 24 hour incubation) would be the best P sinks and that specific marshes could 

be targeted based on this index. It is important to remember, however, that antecedent 

soils of restored wetlands have a finite P retention capacity. The long-term sustainable 

capacity of these systems to retain P is determined primarily by the accumulation of P 

associated with the formation and accretion of new sediments and soils. Studies of 

wetlands constructed to intercept non point source nutrient loads in the MARB confirm 

the importance of sediment accretion for P retention (Anderson et al., 2005; Mitsch et al. 

2005b) but also demonstrate that wetlands can become a P source if sediments are 

remobilized (Mitsch et al., 2005b). Most of the MARB studies represent recently 
161

constructed wetlands, and the long-term sustainable capacity of these systems to reduce P 

loadings is unclear. 

 Wetlands created, enhanced, and restored for N removal could also function for P 

removal, but limits to sustainable P removal must be recognized. Both NO3 and P 

removal in wetlands will be enhanced by longer retention times and accretion of organic 

rich sediments. Long-term solutions for P load reduction in the MARB will likely 

depend more on reduction in sources than will long-term N load reduction. It will be 

important to manage restored wetlands so they do not become long-term sources of P 

after non-point sources of P have been reduced. 

Key Findings and Recommendations 

 As concluded in the Integrated Assessment, wetlands can be very effective in NO3 

removal. Recent data, though limited, support the Integrated Assessment’s conclusion 

that N2O evolution from wetlands restored as NO3 sinks would be a low percentage of 

total denitrification. Wetlands receiving significant non-point source NO3 loads at 

moderate to high NO3 concentrations export comparatively small amounts of organic N 

and are nearly as effective in reduction of total N as in reduction of NO3. This situation 

is less true for wetlands receiving loads at low NO3 concentrations. Hydraulic loading 

rate and NO3 concentration are especially important determinants of NO3 removal rates in 

Corn Belt wetlands. Additional information is needed on created, restored, and enhanced 

wetlands including long-term monitoring for total N and P retention. Based on these 

findings, the SAB Panel offers the following recommendations. 

• Wetland restoration should be evaluated for its full range of benefits. 

• For greatest basin wide reduction in nitrate load, wetland restorations should be 

targeted in those extensively row-cropped and tile-drained areas of the Corn Belt 

where nitrate concentrations and loading rates are highest and sized based on 

expected hydraulic loading rates and load reduction goals. For these areas, nitrate 

mass removal rates could be several times higher than previously predicted. 

• Although limits to sustainable P removal by wetlands must be recognized, 

wetlands restored for N removal should be managed for P retention as well. 

4.5.3. Conservation Buffers 

 Conservation buffer practices include riparian buffers (forests and herbaceous 

cover), field borders, filter strips, contour buffer strips, grass waterways, windbreaks, 

hedgerows, and other practices. They are part of the suite of conservation practices that 

are applied by farmers to achieve productivity, stewardship, and environmental quality 

goals. Conservation buffers differ from other conservation practices in that they will 

require long-term set aside of critical lands from continued agricultural production. 
162

Although often installed under the Conservation Reserve Program (CRP), conservation 

buffers differ from other uses of CRP because conservation buffers allow most land to 

remain in production while using critical areas as buffers for the agricultural land. 

 Prior analysis of nutrient control in the MARB focused on riparian forest buffers, 

one prominent type of conservation buffer (Mitsch et al., 1999). Studies conducted over 

the past decade in the Corn Belt have shown conservation buffers, especially riparian 

forest buffers and riparian herbaceous buffers, to be effective sinks for nutrients and 

sediment in landscapes with a significant portion of water moving as either surface runoff 

or shallow subsurface flow. If nitrate is transported from crop land primarily in tile drain 

flow as in much of the Corn Belt, riparian buffers and vegetated filter strips will have 

little opportunity to intercept nitrate loads. It is likely that if drainage management is 

changed to limit subsurface discharge through tile drains with concomitant increases in 

surface runoff and shallow water table flow, riparian buffers will be critical to achieve 

water quality goals. 

 Reduction of nitrogen by riparian buffers is generally determined by soil type, 

watershed hydrology (artificial drainage, groundwater flow paths, saturation); and 

subsurface biogeochemistry (organic matter supply, redox conditions) (Mayer et al., 

2006). Control of P depends more on infiltration, surface roughness and runoff retention. 

Many riparian buffers have been restored or established, but few have been studied to 

quantify water quality benefits. Richard Schultz, Tom Isenhart and others developed the 

Riparian Management System for application in areas of the Corn Belt dominated by tiledrain systems. Modifications to the original USDA Riparian Buffer specification 

included integration of wetlands to intercept and remove tile drainage nitrate. Lee et al. 

(2000, 2003) reported rates of nutrient and sediment removal in multi-species buffer 

strips intercepting surface runoff in these systems. They found that switch grass and 

switch grass/woody buffers retained 50-80 % of total N, 41 to 92% of NO3-N, 46-93% of 

Total P, and 28-85% of dissolved reactive P from surface runoff produced in simulated 

rainfall events. 

 Riparian herbaceous cover helps reduce sediment and other pollutants in surface 

runoff through the combined processes of deposition, infiltration, and dilution. Those 

functions are due to the cascading influence of perennial vegetation on soil quality when 

compared to soils under annual row-crops. A series of studies on Bear Creek compared 

soil quality and related processes within riparian soils in a corn-soybean rotation with 

those soils in which perennial herbaceous vegetation had been reestablished (Schultz et 

al., 2004). Six years after establishment of riparian switch grass, those soils contained 

more than eight times the belowground biomass as adjacent crop fields (Tufekcioglu et 

al., 2003). As a result, soils in riparian herbaceous cover amassed up to 66 % more total 

organic carbon in the top 50 cm (20 in) than crop-field soils (Marquez et al., 1999). This 

resulted in a two-and-a-half-fold increase in microbial biomass and a four-fold increase in 

denitrification in the surface 50 cm (20 in) of soil when compared to crop-field soils of 

the same mapping unit. As a result of increased soil quality, infiltration was nearly five 

times faster in soils under perennial vegetation than in row-cropped fields (Bharati et al., 

2002). Riparian Management Systems such as those on Bear Creek are well-suited to 
163

intercept increased overland flow that might be associated with changes in drainage 

management. 

 Several researchers have investigated the combined effects of these processes 

within riparian herbaceous vegetation and reported that sediment and nutrients in surface 

runoff can be reduced in the range of 12 to 90 % compared to unbuffered crop fields 

(Dosskey, 2001; Lee et al., 2003). Major differences in impacts on the soil ecosystem 

depend upon the photosynthetic pathway of the dominant vegetation [e.g., C3 (coolseason grasses) or C4 (warm-season grasses)] in a buffer. Riparian herbaceous cover can 

help improve the quality of shallow groundwater, much like filter strips or riparian forest 

buffers. Hydrogeologic setting, specifically the direction of groundwater flow and the 

position of the water table in thin sand aquifers underlying the buffers, generally is the 

most important factor determining buffer efficiency (Dosskey, 2001). 

 When applied as part of a conservation management system, the effectiveness of 

conservation buffers can be enhanced. There are few data on the field or landscape level 

effectiveness of conservation buffers applied with or without other conservation 

measures. Most data are from plot studies. Plot studies are inadequate, especially for 

studies of grass waterways (GWW), which are designed to convey overland flow from 

fields and stream bank restoration designed to reduce loss of sediment and sediment 

bound chemical from unstable banks. Because GWW are installed in areas of known 

water flow, they avoid problems of runoff bypassing filter strips and field borders. The 

few studies of GWW conducted at the field scale show that they are very effective at both 

runoff reduction and sediment trapping. In Germany, unmanaged grass waterways 

reduced runoff and sediment delivery by 90 and 97% respectively compared to adjacent 

fields with no GWW (Feiner and Auerswald, 2003). A GWW that was mowed closely 

was less effective, with reductions of 10 and 27% for runoff and sediment delivery, 

respectively. In New Brunswick, Canada, Chow et al. (1999) compared up- and downslope cultivation of potatoes and grain to the same crops with a terrace and grass 

waterway system. The conservation system reduced runoff by 31% and sediment 

delivery by 78%. On three small watersheds in the claypan soils region of Missouri, 

sediment and TP loss increased as the extent of GWW decreased (Udawatta et al., 2004). 

 There are ongoing efforts by USDA to estimate the impacts of conservation 

buffers on water quality in all watersheds with significant amounts of agriculture. The 

Conservation Effects Assessment Project (CEAP) will eventually provide model-based 

estimates of the water quality impacts of conservation practices in the MARB (Kellogg 

and Bridgham, 2003). Conservation buffers are an important component of USDA 

conservation programs. Table 12 summarizes the extent of seven major conservation 

buffer practices installed in the six sub-basins of the MARB in federal fiscal years 2000 

through 2006 (October 1999 - October 2006) (M. Sullivan, personal communication, 

based on USDA-NRCS-Performance Results System, 

http://ias.sc.egov.usda.gov/prshome). An estimated 0.94 million ha (2.31 million ac) of 

conservation buffers were installed in the MARB in 1999-2006. As shown, each ha of 

conservation buffer treats one or three ha of adjacent agricultural land, giving an 
164

estimated 3.46 million ha (8.55 million ac) of agricultural land has been treated by these 

six conservation buffer practices (Table 12). 

Table 12: Areas (ha) of conservation buffers installed in the six sub-basins of the MARB for FY 2000 - 

FY2006. 

Subbasin Contour 

Buffer 

Strips 

(ha) 

Field 

Border 

(ha) 

Filter Strip 

(ha) 

Grassed 

Waterway 

(ha) 

Riparian 

Forest 

Buffer (ha) 

Stream 

bank 

Protection 

(km) 

Windbreaks 

and 

Shelterbelts 

(ha) 

Conservat

ion 

Buffers 

Applied 

(ha) 

Ohio 3,362 5,441 50,617 21,346 32,497 755 794 114,832

Tennessee 196 1,914 10,724 817 10,752 418 2 26,025

Upper 

Mississippi 

22,217 7,357 159,604 43,421 75,139 722 8,448 317,422

Lower 

Mississippi 

165 7,541 10,274 661 56,106 503 391 75,486

Missouri 7,374 16,413 116,755 31,067 31,492 470 39,377 256,693

Arkansas 

White-Red 

1,883 15,631 79,658 8,197 29,745 287 2,173 145,290

Sum 35,196 54,298 427,631 105,507 235,731 3,155 51,185 935,748

Area 

treated 

(ratio) 

1:1 1:1 3:1 3:1 3:1 NA 3:1 

Area 

treated 

70,393 108,595 1,710,52

5

422,030 942,926 NA 204,739 3,459,207

* Kilometers are shown for stream bank protection. Conservation buffers applied includes areas in other practices not 

shown here that are cumulatively small areas compared to the practices shown. The areas treated are based on the 

ratios shown and assumes that each ha of buffer treats either one ha or three ha of adjacent agricultural land. Areas of 

practices are from Mike Sullivan, USDA-NRCS, Personal Communication, and are derived from NRCS-PRS, 

http://ias.sc.egov.usda.gov/prshome. 

 Information on the extent of other conservation practices established from FY 

2000 through FY 2006 is also available from the NRCS Performance Results System. 

Practices that are applied each year such as conservation tillage, residue management, 

and nutrient management may be reported more than once during the record period if 

there is a change in owner/operator, a new conservation plan is developed, and associated 

practices are reported. There may have also been some systematic annual reporting in the 

early years of the record period (2000-2003) (Personal communication, Mike Sullivan, 

USDA-NRCS). All conservation tillage and residue management practices combined 

were applied on as much as 8.42 million ha (20.8 million ac) and nutrient management 

was applied on as much as 7.4 million ha (18.3 million ac) in the MARB in FY 2000 to 

FY 2006 (Mike Sullivan, USDA-NRCS, Personal Communication, based on NRCSPRS). Wetland creation, enhancement and restoration was applied on 0.57 million ha 

(1.42 million ac), drainage water management was applied on 756 ha (1,867 ac), and 

stream bank restoration was installed on 3,155 km (1,972 mi). The values for 2002-2005 

were reported in the USEPA Management Action Review Team report (MART 2006a) 

and are similar to the above numbers when put on the same year basis. 
165

 Currently no national databases allow a more detailed estimation of the 

environmental benefits of these conservation practices, including conservation buffers. 

This is the goal of the CEAP project. Estimates can be made based on acreage values, 

but these cannot take into account either placement or efficacy of practices. 

Cumulatively, conservation buffers, residue management, nutrient management, and 

wetlands have impacted up to 21 million ha (51.9 million acres) of agricultural land in the 

MARB based on the FY 2000 - FY 2006 areas of conservation practices. This area is the 

sum of residue management, nutrient management, conservation buffer acreage, wetland 

acreage and the potential land treated by conservation buffers (Table 12) and wetlands 

(assuming 3 hectares treated for 1 hectare of wetlands). In reality, conservation practices 

are applied as a system of practices, and it is likely that the total area treated through 

these practices is less than 21 million ha (51.9 million ac). Additionally the data bases 

used are likely to include some duplicate reporting for the annual practices. The nutrient 

load reductions for these practices could be estimated based on amounts of N and P load 

retained. Although these would be crude estimates, they would provide numbers for 

comparison to the nutrient load reduction goals and provide a rough idea of where 

conservation programs stand relative to those goals. 

Key Findings and Recommendations 

 Conservation buffers and other conservation practices have affected a significant 

acreage of MARB cropland through existing federal, state, and private programs. The 

SAB Panel offers the following recommendations. 

• Continued, new, and enhanced small watershed based studies of suites of 

conservation practices as applied on farms and in agricultural watersheds are 

necessary. Analysis of effects of conservation buffers and other conservation 

practices in the MARB should be coordinated with the ongoing USDA 

Conservation Effects Assessment Project. 

• Conservation buffers and other conservation practices in the MARB should be refocused on N and P retention with special attention given to the interactions of 

buffers with other practices. Environmental benefits indices should be calculated 

in a way as to provide extra weight for N and P retention. 

4.5.4. Cropping systems 

 Current cropping systems within the MARB are well established, but advances in 

N fertilizer production technology, innovative crop rotations, inter-seeding with cover 

crops, and alternative mulches or crop residues provide opportunities to improve water 

and nutrient use efficiency as well as to decrease leaching and runoff of nutrients and 

sediments. For example, inter-seeding of a leguminous cover crop within existing crop 

rotations could enhance N and P use efficiencies, as long as the cover crop is carefully 
166

managed. Also, greater adoption of perennial systems, which could include cellulosic 

production, have the potential to influence nutrient export via reduced N and P 

applications as well as altered water budgets. Evapo-transpiration and infiltration will 

likely be greater with perennial than annual cropping systems, contributing to a decrease 

in potential runoff. Hydrologic and water quality issues related to perennials and 

cellulosic production are discussed in more detail in Section 4.5.9. – Ethanol and Water 

Quality in the MARB. 

 A continuous corn rotation typically results in annual N fertilizer applications 

between 150 and 250 kg N/ha (134 and 223 lb N/ac). This is a large amount of N 

fertilizer relative to amounts applied to other crops. Including other crops (particularly 

legumes) in a crop rotation usually reduces annual N fertilizer applications needed. In 

addition to applying less N, perennial crops, such as alfalfa or other grass mixtures, have 

longer effective growing seasons and are more efficient N users than annual crops, which 

translate to greater water use and less nitrate leaching. 

 Randall et al. (1997) compared tile drainage and nitrate loss for corn-soybean and 

corn-corn rotations to alfalfa and Conservation Reserve Program (CRP) grassland. From 

770 to 905 mm (30 to 36 in) of tile water was recorded for the corn-corn and cornsoybean rotations from1988-1993, whereas 416 to 640 mm (16 to 25 in) of tile water was 

recorded for alfalfa and CRP. Flow-weighted nitrate-N concentrations were less than 5 

mg/L for alfalfa and CRP but ranged between 13 and 40 mg/L for the rotations including 

corn and soybean. The four-year nitrate-N loss from continuous corn or corn-soybean 

rotations was 202 to 217 kg N/ha (180 to 194 lb N/ac), while for alfalfa and CRP the loss 

was less than 7 kg N/ha (6 lb N/ac). Similarly, Jaynes et al. (2001), showed for a cornsoybean rotation in central Iowa that even at economically optimum N fertilizer rates for 

corn (67 to 172 kg N/ha or 60 to 154 lb N/ac), NO3 loss in tile drainage water increased 

from 29 to 43 kg N/ha (26 to 38 lb N/ac) with application rate. Also, a net N mass 

balance indicated that N was being mined from the soil at economically optimum N 

fertilizer rates and the system would not be sustainable (Jaynes et al., 2001). 

 Besides crop selection to enhance N and P removal, crop rotation also can be 

managed to maximize nutrient removal and minimize leaching. Together, crop selection 

and rotation can influence the amount of N and P in a soil profile as well as water 

available for nutrient leaching. As mentioned, legumes, such as alfalfa and soybean, that 

do not require supplemental N, can effectively use or "scavenge" residual inorganic N 

remaining in the soil from previous crops. Some crops take up more P, and deep-rooted 

crops can remove N and P from subsoil horizons. For example, root development of a 

typical 3-year continuous corn system (maximum depths in May through September) 

does not always coincide with time of high NO3 leaching potential (generally February to 

April). An alternative cropping system comprised of corn-winter wheat-alfalfa provided 

a much different root development pattern, one that should more efficiently retain N 

because it has deeper roots that are present most of the year (Sharpley et al., 2006b). 

Olson et al. (1970) found that NO3 concentrations at a depth of 1.2 to 1.5 m (3.9 to 4.9 ft) 

in a silt loam soil were lower for an oat-meadow-alfalfa-corn rotation than for continuous 

corn when ammonium nitrate was applied to both systems. The reduction in NO3
167

leaching was directly proportional to the number of years that oats, meadow, or alfalfa 

was grown in rotation with corn. The reduction was attributed to the combined recovery 

of NO3 by shallow-rooted oats, followed by deep-rooted alfalfa (Olson et al., 1970). The 

potential for NO3 leaching in such rotations is, therefore, less when compared with 

continuous annual monocropping systems. 

 Clearly, including perennial crops in a rotation, as well as conversion to perennial 

systems, can reduce NO3 leaching, partly due to the fact that perennials are generally 

more efficient users of N than annuals. As a result, Randall and Vetsch (2005) raises a 

key question of whether significant reductions in nutrient (especially NO3) loadings to 

surface waters are possible without changing from the predominant annual cropping 

system of corn-soybean rotation to a mixed system that includes perennials. While 

annual grain crop production is an essential component of agricultural systems in several 

areas of the MARB, the development of economically viable continuous cropping 

systems will help improve in-field nutrient use efficiency and decrease off-site loads. 

Additional co-benefits of perennials such as switchgrass, are that they have the potential 

to accumulate large amounts of below-ground biomass and are effective in sequestering 

C (McLaughlin and Walsh, 1998; McLaughlin and Lszos, 2005). 

Retirement of land through the Conservation Reserve Program has demonstrated 

different results for various cropping systems. For lands previously in corn, the reduction 

in N delivered to the Mississippi River may have been as much as 25 to 30 kg N/ha/yr 

(22 to 27 lb N/ac/yr). For soybean it would have been somewhat less, and for small 

grains, particularly wheat in the High Plains, smaller reductions, in the range of 10 kg 

N/ha/yr (8.9 lb N/ac/yr) may have been realized (see Section 3.1.2. – Subbasin Annual 

and Seasonal Flux). Where CRP has been used to establish buffers, not only are 

reductions from the retired lands realized, but the buffers can also be effective in 

reducing inputs of N and P from upslope cropland entering water courses via surface 

runoff and shallow subsurface flow. It should be noted, however, that most land enrolled 

in CRP is primarily sloping, erosive land that is not tile drained. For instance, McIsaac 

and Hu (2004) studying N flux in several Illinois rivers between 1977 and 1997 found 

that riverine N flux was about 100% of net N input for the tiled drained region (27 kg 

N/ha/yr or 24 lb N/ac/yr). In the non- tile drained region, riverine N flux was between 25 

and 37% of net N input (23 kg N/ha/yr or 20 lb N/ac/yr). 

Key Findings and Recommendations

 Cover crops and other living mulches can improve water and nutrient use 

efficiencies and reduce nitrate leaching. Further research and demonstration is needed in 

the MARB in several areas: examining the benefits of intercropping cover crops with 

annuals such as corn; determining if leguminous cover crops reduce fertilizer N 

requirements; and assessing how changes in cropping patterns can impact nutrient loss at 

both local and basin-wide scales. If farmers could be encouraged to switch to a rotation 

of perennial crops as compared to the predominant corn-soybean rotation system, 
168

significant N and P reductions would result. Based on these findings, the SAB Panel 

offers these recommendations. 

• Cover, relay, and perennial crops should be considered in alternative cropping 

systems that will reduce nutrient loss. Cropping systems that efficiently include 

cover crops in grain and row cropping should also be encouraged in the Corn Belt 

region of MARB. This should focus on the use of fall planted small grain cover 

crops more suited to the short growing season after harvest and cold winters of the 

upper Midwest. 

• Where corn-soybean production systems exist and/or where it is not feasible to 

plant cover crops, it is even more important to encourage off-field conservation 

practices. 

4.5.5. Animal Production Systems 

System development and nutrient flows 

 While overall production livestock numbers in the MARB have declined (see 

Section 3.2. – Mass Balance of Nutrients), there has been an intensification of operations 

in certain areas (see Figure 43, Figure 44, and Appendix E: Animal Production Systems). 

Farmers adopted the AFO paradigm because of competitive pressures, changing 

marketing practices, a need to be responsive to consumer demand for quality meat 

products at a low cost, and declines in income from traditional grain crops in certain areas 

of the MARB with inherently infertile soils (Lanyon, 2005). This critical socioeconomic 

shift must be considered when proposing changes within the MARB that decrease the 

impact of AFO and manure management on nutrient export. 
169

Figure 43: Recoverable manure N, assuming no export of manure from the farm, using 1997 census data. 

Adapted from USDA (2003) with the author’s permission. 

 As a consequence of the spatial separation of crop and animal production systems, 

fertilizer N and P is imported to areas of grain production. The grain (harvested N and P) 

is then transported to areas of animal production, where inefficient animal utilization of 

nutrients in feed (less than 30% is utilized) are excreted as manure. This system has led 

to a large-scale, one-way transfer of nutrients from grain- to animal-producing areas 

within the MARB and dramatically broadened the emphasis of nutrient and manure 

management strategies from field to watershed to basin scales. For the MARB, farmlevel nutrient excesses are estimated at 337 million kg N (743 million lb N) and 242 

million kg P (534 million lb P) (Gollehon et al., 2001). 
170

Figure 44: Recoverable manure P, assuming no export of manure from the farm, using 1997 census data. 

Adapted from USDA (2003) with the author’s permission. 

 The land application and discharge of nutrients in manure from AFOs are 

regulated under the National Pollutant Discharge Elimination System (NPDES), which 

generally define an AFO as an operation where livestock are confined for an extended 

period of time (at least 45 days in a 12-month period) and there’s no grass or other 

vegetation in the confinement area during the normal growing season (U.S. EPA, 2000a). 

This definition is intended to differentiate confinement-based operations from pasturebased operations, which are excluded from the Confined Animal Feeding Operations 

(CAFO) regulations. The NPDES permit is required to control pollutants at an AFO and 

keep them from entering surface waters. More explicitly, the U.S. EPA (2000a) defines 

CAFOs as livestock operations that meet one of the following characteristics: 

• Confine more than 1,000 animal units (AU), where 1,000 AUs are defined as 

1,000 slaughter and feeder cattle, 700 mature dairy cows, 2,500 swine (other than 

feeder pigs), 30,000 laying hens or broilers if the facility uses a liquid system, and 

100,000 laying hens or broilers if the facility uses continuous overflow watering. 

• Confine between 300 and 1,000 AU (as defined above), and either a man-made 

ditch or pipe carries manure or wastewater from the operation to surface water or 

animals come into contact with or surface water running through the area where 

they are confined. 

These regulations are enacted at a national level, and thus, there are recommendations 

and controls on the land application or utilization of manures and their component 

nutrients are in place at a state level in the MARB. Based on an US EPA summary of 
171

CAFO permit implementation completed in the first quarter of 2007, less than half of the 

CAFOs in the MARB were permitted (46%; Table 13). States included are in Table 13, 

if part of the state drains into the MARB. The approximate number of permitted CAFOs 

in the MARB is similar to the national average (44%; U.S. EPA, 2007) but clearly, rule 

implementation varies among states. 

Table 13: Status of implementation of permits under the 2003 CAFO rule for states within the MARB. 

Data provided by EPA Office of Wastewater Management, 2007. 

State Number of 

CAFOs 

Number of CAFOs 

with permits to date 

Permit coverage for 

CAFOs under 2003 rule 

Alabama 558 440 79 

Arkansas 2,110 70 3 

Colorado 225 33 15 

Illinois 500 8 2 

Indiana 584 413 71 

Iowa 1,859 113 6 

Kansas 476 462 97 

Kentucky 150 67 45 

Louisiana 150 2 1 

Michigan 198 56 28 

Minnesota 1,007 1,000 99 

Mississippi 433 190 44 

Missouri 492 492 100 

Montana TBD 75 TBD 

Nebraska 1,000 303 30 

New Mexico 151 47 31 

North Carolina 1,222 1,200 98 

North Dakota 47 0 0 

Ohio 162 64 40 

Oklahoma 625 163 26 

Pennsylvania 462 165 36 

South Dakota 369 303 82 

Tennessee 129 130 101 

Texas 1,204 639 53 

Virginia 150 0 0 

West Virginia 30 0 0 

Wisconsin 161 161 100 

Wyoming 51 47 92 

Total 14,505 6,643 46 
172

Manure as a Component of N and P Mass Balances 

 Within the MARB, counties with the greatest excess of recoverable manure N and 

P (if applied on the farm where it is generated) tend to be in the western and drier areas of 

the basin, Arkansas, and central Minnesota (Figure 43 and Figure 44). Recoverable 

manure is defined as the portion of manure as excreted that could be collected from 

buildings and lots where livestock are held and, thus, would be available for land 

application. Recoverable manure nutrients are the amounts of manure N and P that 

would be expected to be available for land application (USDA, 2003). They are 

estimated by adjusting the quantity of recoverable manure for nutrient loss during 

collection, transfer, storage, and treatment. Recoverable manure nutrients are not

adjusted for losses of nutrients at the time of land application. Where riverine N export is 

the greatest (upper Mississippi and Ohio River basins with tile drainage), manure N 

excess tends to be less. lower Mississippi River basin states, particularly Arkansas and 

northern Missouri, clearly have more manure P on some farms than land area to apply it 

(Figure 43). Although N from manure can be important in specific areas, basin-wide N 

loss is a result of the dominant inputs of fertilizer and N2 fixation on tile-drained corn 

and soybean fields. For P, manure is a more important source, particularly on the western 

side of the basin (Figure 44). 

 Large-scale consolidation has created much larger AFOs, which makes 

economical utilization and re-distribution of manure to croplands difficult and has 

profound consequences for regional nutrient transfer and management within the MARB. 

Furthermore, the potential for co-locating AFOs with areas of the corn production for 

ethanol generation may exacerbate the accumulation of manure-based nutrients in these 

areas. This co-location stems from the use of by-products from ethanol production 

(distiller’s grain) as animal feed (for more information see Section 4.5.9). 

Remedial Strategies 

 Manure is a valuable resource for improving soil structure and increasing 

vegetative cover, thereby improving water quality via reduced runoff and erosion 

potential. Manures have been historically applied at rates designed to meet crop N 

requirements. This has resulted in the accumulation of soil P above levels required for 

crop production, and a concomitant increase in the potential for N and P loss via runoff, 

leaching and N2O emission within the MARB (Table 14; Aillery et al., 2005; Sharpley et 

al., 1998). In the past, separate strategies for either N or P have been developed and 

implemented at farm or watershed scales. The SAB Panel recognizes that this approach 

needs to change; N and P need to be managed jointly in order to improve water quality. 

Because of different critical sources, pathways, and sinks controlling N and P export, 

remedial strategies directed at only N or only P control can negatively impact the other 

nutrient. For example, basing manure application on crop N requirements to minimize 

nitrate leaching can increase soil P and enhance P losses (Sharpley et al., 1998; Sims, 

1997). In contrast, reducing surface runoff losses of P via conservation tillage can 

enhance nitrate leaching in some cases (Sharpley and Smith, 1994). 
173

Table 14: Estimates of manure production and N and P loss to water and air from Animal Feeding 

Operations within the Mississippi River basin, on information from the 2002 U.S. Census of Agriculture 

(adapted from Aillery et al., 2005). 

Region of MARB # 

Operations 

Total 

Manure 

N 

Runoff 

N 

Leached 

N 

Emissions 

Total 

N loss 

P 

Runoff 

 million Mg - - - - - - - - - - - - - - million kg - - - - - - - - - - - - - - 

Lake States (MI, 

MN, WI) 52,498 62.52 32.89 0.36 164.45 198 5.58 

Corn Belt (IA. IL, 

IN, MO, OH) 71,252 85.09 39.73 0.47 234.89 275 11.78 

Northern Plains 

(KS, ND, NE, SD) 26,087 72.27 36.31 0.37 168.44 205 6.99 

Appalachia (KY, 

NC, TN, VA, WV) 22,776 79.57 54.65 0.91 259.16 315 15.79 

Delta States (AR, 

LA, MS) 12,252 19.97 8.92 0.15 62.57 72 4.47 

Southern Plains 

(OK, TX) 10,500 49.19 21.96 0.20 119.74 142 7.72 

Total 195,365 368.63 194.46 2.46 1009.26 1206 52.34 

 

 Long-term sustainable management of nutrients in manure begins with sound feed 

decisions, which generally lie with the integrator in the CAFO industry rather than the 

individual farmer. Nutrient inputs to a farm should be matched as closely as possible 

with export as animal or crop products. If a farm’s N and P budget is rich in imports, 

regardless of any other management decisions, there will be an ongoing accumulation of 

N and P on the farm, which in the long-term will ultimately increase the potential for 

nutrient loss to water or air when manure is land-applied. Nevertheless, the short-term 

impacts of land-applying manure or litter on nutrient loss can be reduced by the adoption 

of conservation practices detailed by USDA-NRCS (ftp://ftpfc.sc.egov.usda.gov/NHQ/practice-standards/standards/590.pdf). However, conservation 

measures at both farm and watershed scales involves a complex suite of options, which 

must be customized to meet site-specific needs (for more information see Section 4.5.10: 

Integrating Conservation Options and Appendix E: Animal Production Systems). 

Alternative manure management technologies 

 Reducing farm-gate inputs of N and P in animal feed presents one of the best 

nutrient management opportunities to effect a lasting reduction in N and P loss 

(Appendix E: Animal Production Systems). Other measures, generally aimed at 
174

reducing the potential for N and P losses, are seen as short- rather than long-term 

solutions to environmental concerns. For instance, long-term monitoring of P budgets in 

Ohio showed that after nearly 20 years of BMP adoption and despite continually 

increasing soil test P levels, manure applications and timing have been managed better, 

resulting in more efficient use of P and reduced P loss to surface waters (Baker and 

Richards, 2002). Manure-related conservation practices include: 

• Manure amendments, such as alum, to reduce ammonia volatilization and 

sequester P in less soluble forms; 

• coagulant and flocculent techniques to separate and concentrate nutrients in liquid 

manure systems; and 

• combining manure with biosolids and woodchips to reclaim soils that have been 

disturbed (e.g., by mining or urban development). 

 As the cost of N fertilizer increases, it is clear that new markets for alternative 

uses or products for manure will open up. For example, on-farm and regional energy 

production via burning of manure is of increasing cost-effectiveness. Ash production via 

burning, while rich in P, will be appreciably less bulky and, thus, enable cost-effective 

transportation further from the source of generation. The bulky nature of manures and 

resulting high cost of transportation has always been a major limitation to more effective 

redistribution of N and P to nutrient deficient areas of the MARB. 

 Recent efforts to exclude cattle from streams as part of the Conservation Reserve 

Enhancement Program (CREP) were estimated to have resulted in a 32% decrease in P 

loadings to streams within the Cannonsville watersheds in south central, New York 

(James et al., 2007). Thus, exclusionary programs like CREP and stream bank fencing 

are working to reduce nutrient loading by fencing cattle out of the stream and adjacent 

riparian zones. Clearly, grazing management and placement of stream bank fencing is 

important to minimizing watershed export of P. For instance, herd size, pasturing time, 

and cattle type could all be used to prioritize sites for stream bank fencing installation. In 

addition, field observations [such as those by James et al. (2007)] show installation of 

alternative watering sources do not necessarily preclude continued use of streams as a 

preferred water source. 

 The wider adoption of manure hauling that links producers with buyers will 

greatly enhance the sustainability of AFOs. At a state level, the Discovery Farms 

program is conducting research on privately-owned Wisconsin farms in different 

geographic areas, facing different environmental challenges (see 

http://www.uwdiscoveryfarms.org/new/index.htm). The Discovery Farms program has 

been very successful at gaining farmer support in at-risk catchments in efforts to find the 

most economical solutions to overcoming the challenges environmental regulations place 

on farmers. At a watershed level, the Illinois River Watershed Partnership (see 

http://www.irwp.org/index.html) was established in 2005 to improve and protect water 
175

quality in the Illinois River in Arkansas and Oklahoma by working at a grassroots level 

with watershed citizens and other organizations. 

Key Findings and Recommendations 

 The impacts of animal production systems are mainly expressed at a local rather 

than MARB scale. Overall, numbers of animals in the MARB have decreased, but 

localized increases have occurred in several regions, which have had an impact on local 

water resources. The economic and environmental sustainability of AFOs hinges on 

reducing the nutrient imbalance at farm and watershed scales through carefully managed 

feeding strategies. The wider adoption of manure transportation that links producers with 

buyers will greatly enhance the sustainability of AFOs. The large-scale consolidation of 

AFOs, co-siting with biofuel production facilities (byproduct grains used as animal feed), 

and increases in N fertilizer prices will likely create the economies of scale and 

alternative technologies for on-farm or localized manure use and management more 

feasible. 

The success of non-profit programs supported by watershed agricultural councils, 

industry, and state agencies, should provide valuable demonstration models. If energy 

prices remain at current levels, bioenergy production from manures could provide an offfarm market for manures and reduce localized nutrient surpluses. Continuing educational 

efforts with farmers and the public regarding the importance and impact of conservation 

practices will be essential to reach environmental goals. Based on these findings, the 

SAB Panel offers the recommendations below. 

• Strategies need to be implemented to encourage further development of 

alternative uses for manures such as in composting, pelletizing, and granulation, 

and as a soil amendment in nutrient deficient areas of the MARB. 

• Land-management planning and implementation of conservation practices should 

be designed to identify and avoid applications in critical loss areas, to use buffers 

or riparian zones, to manage grazing, to exclude stream banks, and to use 

subsurface injection with innovative applicators. 

• Incentives to encourage on-farm and local bioenergy production from manure 

sources should be provided. 

4.5.6. In-field Nutrient Management 

Fertilizer sources

 The principal fertilizer N sources (>90% of fertilizer N) used in the MARB are 

anhydrous ammonia, urea-ammonium nitrate solutions, and urea. Anhydrous ammonia 

use in several leading corn-producing states (IL, IN,IA, MN, NE, OH) has tended to 

decline in recent years, perhaps with the exception of consumption in Illinois and Indiana 

(Figure 45) (Sources: Association of American Plant Food Control Officials; H. Vroomen 
176

with TFI-personal communication, 2007). The largest decline has been in Nebraska, 

where use of anhydrous ammonia N has declined about 40% since the mid-1980s. 

1970 1975 1980 1985 1990 1995 2000 2005 2010

Anhydrous Ammonia

(metric tons N)

0

100000

200000

300000

400000

500000

600000

700000

Illinois

Indiana

Minnesota

Iowa

Ohio

Nebraska 

Figure 45: Fertilizer N consumption as anhydrous ammonia in leading corn-producing states for years 

ending June 30. 

The combined N consumption of urea and urea-ammonium nitrate solution has increased 

and recently surpassed anhydrous ammonia tonnage in these six leading corn-producing 

states (Figure 46). Although these data illustrate shifts in fertilizer N sources used, they 

do not allow conclusions about the portion of the annual anhydrous ammonia 

consumption that may be applied in the fall. 
177

0

500,000

1,000,000

1,500,000

2,000,000

2,500,000

3,000,000

3,500,000

4,000,000

1965 1970 1975 1980 1985 1990 1995 2000 2005 2010

Year

Metric tons of N

Anhydrous ammonia 

Urea + UAN Sol.

Anhdrous ammonia + urea + UAN

Figure 46: Changes in the consumption of principal fertilizer N sources used in the six leading cornproducing states (IA, IL, IN, MN, NE, and OH) for years ending June 30. 

Fertilizer use and application technology 

 The Integrated Assessment (CENR, 2000) concluded that “discharges of nitrogen 

from farms to streams and rivers could be reduced by implementing a wide variety of 

changes in management practices”. These practices include switching from fall fertilizer 

N to spring N applications, and applying nitrogen fertilizer and manure at not more than 

agronomically recommended rates. Application rate and timing are linked for N because 

the closer application is to the time of crop need, less N is lost to the atmosphere and 

water, and less N is needed. Research at five Management System Evaluation Area 

(MSEA) sites in the MARB (OH, IA, MN, MS, NE) reaffirmed BMPs for water quality, 

including soil nitrate tests, improved water management, and improved N timing and 

placement relative to crop needs (Power et al., 2000). Determining N sufficiency by 

monitoring for plant greenness and use of field or remote-sensing technologies followed 

by site-specific N applications hold promise to manage N more precisely. 

Application timing. The risk of N loss with corn is greatest when fertilizer is 

applied some time before the period of rapid plant growth. Data on fall application are 

not directly available for the MARB and even seasonal data on fertilizer sales are not 

kept by all states in the MARB (Terry, 2006). Fertilizer sales records for Iowa (from July 

2002 to June 2006) showed that 48% of N fertilizer was sold in the period from July to 

December and 52% from January to June. For anhydrous ammonia, the most common N 

form used and the primary form applied in the fall, 54% was sold in the period from July 

to December and 46% was sold from January to June. July to December sales of 
178

anhydrous ammonia accounted for 273,000 tons of actual N (data from 

http://www.agriculture.state.ia.us/fertilizerDistributionReport.htm). For Illinois, there 

has been an increase in fall N sales from the 1970s and 1980s to present, from about 25% 

to 40 to 50% (Figure 47). 

 Although it is not possible to correlate fall N application directly with fall 

fertilizer N sales, it is likely that a large fraction of the fall N sales represents N applied in 

the fall (Czapar et al., 2007). A portion of this fall N tonnage sold may also be stored at 

dealerships or in on-farm storage vessels for application the following spring. The 

USDA Agricultural Management Resource Survey (ARMShttp://www.ers.usda.gov/Data/ARMS/app/Crop.aspx ) data provide some insight into fall 

N applications, yet they are not sufficiently complete (i.e. key years are missing) to 

determine if the percentage of the acreage that receives some amount of fall N is 

increasing, decreasing, or remaining static (Figure 47). The data do indicate that 

Minnesota, Iowa, and Illinois tend to fall apply some N on a larger fraction of their corn 

acres, compared to the other three states shown in (Figure 47). For three states, USDA 

ARMS data were used to calculate the total fraction of N applied to corn in the fall, and 

IL sales were also compared (Figure 49). As fewer producers in the Corn Belt farm the 

existing acreage, there has been greater pressure to complete fertilization in the fall, 

because of the numerous logistical challenges (labor demands, transportation and 

application equipment availability, weather uncertainty, and fertilizer supply and cost 

uncertainty) in the spring. 

0

10

20

30

40

50

60

70

80

90

1995 1997 1999 2001 2003 2005 2007

Year

Corn acreage fertilized with N 

which received some amount 

of N in fall, %

IL

IN

IA

MN

NE

OH

Average

Figure 47: Percentage of N fertilized corn acreage which received some amount of N in the fall. 
179

0

10

20

30

40

50

60

ARMS data

IL Sales 

Year

1994 1996 1998 2000 2002 2004 2006 2008

0

10

20

30

40

50

Fraction Fall Fertilizer N Applied (% of total for crop)

0

10

20

30

40

50

Illinois

Iowa

Minnesota

Figure 48: USDA ARMS data for the three states with highest fall N application, showing total amount of 

fall applied N for that crop. Also shown are Illinois sales data for the same period. 

 Randall and Sawyer (2005) contacted State Extension soil fertility specialists and 

State Fertilizer Associations to determine the fertilizer N amount that is applied in the 

fall. Based on these data, they estimated 25% (5.1 million ha or 12.9 million ac) of the 

20.5 million ha (50.6 million ac) of corn in an 8-state area (IA, IL, IN, MI, MO, MN, OH, 

WI) received N in the fall. States with the largest amount of fall-applied N were 

Minnesota (1.85 million ha or 4.56 million ac), Iowa (1.42 million ha or 3.52 million ac), 

and Illinois (1.33 million ha or 3.28 million ac). It is likely that tile-drained portions of 

these eight states have higher proportions of N applied in the fall either because of a 
180

greater dominance of corn/soybean agriculture or because regional soil temperatures are 

also cold enough to help minimize the conversion of ammonium-N to nitrate-N 

(nitrification) in the fall. Fall N application for corn as anhydrous ammonia, is currently 

a recommended practice by virtually all Land Grant universities in the cornbelt, where 

soil temperatures are consistently below 50o

 F at the 1.2 to 1.8 cm depth (0.47 to 0.72 in 

or about 1⁄2 to 3⁄4 in), and the risk of environmental loss is not considered high or a 

pragmatic concern (Snyder et al., 2001). Additional guidance is usually provided in 

publications by Land Grant universities to maximize the benefits of fall N application and 

to help minimize the risk of economical and environmental N losses (e.g. Shapiro et al., 

2003; Bundy, 1998). 

 In a 2003 phone survey of Champaign Co., IL (a dominantly tile-drained area), 

61% of the 352 respondents reported applying some N in the fall, and 49% of 

respondents applied all of their N in the fall (von Holle, 2005). Overall, the farmers who 

fall fertilized applied an average of 79% of their annual N needs before January 1, 2003. 

Data from 11 tile drained central-Illinois counties showed generally greater fall N 

fertilizer sales than the state as a whole (Illinois Department of Agriculture fertilizer 

tonnage reports). This difference was primarily due to the southern and non-tile drained 

portion of the state having winter soil temperatures that are too warm for fall application, 

where it is not recommended. 

Crop Year

1975 1980 1985 1990 1995 2000 2005 2010

Fall Sales (% of total)

10

20

30

40

50

60

70

Illinois

11 tile drained

central-Illinois counties

Figure 49: Fraction of annual fertilizer N tonnage in Illinois sold in the fall. 

 The effects of fall N application versus spring N application on nitrate transport in 

tile drainage depend on many factors, including soil temperatures, soil texture, 

precipitation, and drainage intensity. Randall and Sawyer (2005) reviewed the timing of 
181

N application and determined that spring application in Minnesota will typically result in 

15% less nitrate-N loss than with fall application. In areas with warmer non-growing 

season temperatures (such as central Illinois) that are tile-drained, losses of fall applied N 

may be greater. Watershed-scale studies of changing from fall to spring application 

(side-dressed) and changing the rate to account for more efficient use of spring applied N 

showed at least a 30% reduction of nitrate concentrations in tile drain water (Jaynes et al., 

2004). These studies indicate there is a great potential in some years for substantial 

reductions in N loss by applying N closer to when the crop can utilize it efficiently. 

If these various estimates of N use and nitrate-N loss are combined, changing 

from fall to spring application may affect at least 25% of the corn acreage and reduce 

nitrate-N losses to streams from those acres by perhaps 10 to 30%. Split applications of 

N do not always result in increased N efficiency and reduced nitrate-N losses, just 

because of improved N synchrony with crop uptake demands. The literature to support 

this practice indicates mixed results (Randall and Sawyer, 2005). 

 Nitrification inhibitors delay the conversion of ammonium to nitrate in soil. In 

Illinois, it is estimated that a nitrification inhibitor is added to about 50% of the fall 

applied anhydrous ammonia (Czapar et al., 2007). Application of a nitrification inhibitor 

with anhydrous ammonia in the fall increased apparent recovery of N fertilizer in the corn 

grain from 38% without a nitrification inhibitor to 46% with an inhibitor, compared to 

47% with spring application with no nitrification inhibitor in long-term research results in 

Minnesota (Randall and Sawyer, 2005, Randall et al., 2003). Ferguson et al. (2003) 

found that in Nebraska the benefits of nitrification inhibitors (either increased yield or 

reduced NO3–N leaching) are strongly dependent on specific conditions and are most 

likely to be observed at suboptimal N rates (i.e. <economically optimum N rate (EONR; 

the point where the last increment of N returns a yield increase large enough to pay for 

the additional N)). They also reported that nitrification inhibitors can reduce crop yields 

with late sidedress N applications. It is well known that time of N application will 

largely govern any benefits from the use of nitrification inhibitors. Assuming increased 

N recovery by the crop translates to less nitrate leaching, nitrification inhibitors can 

potentially provide an economic benefit to farmers while reducing leaching. 

 Although the fertilizer N use trends indicate increased urea and urea-ammonium 

nitrate (UAN) solution use in the Cornbelt and lower anhydrous ammonia use (Figure 

46), there is a need for more research to document the benefits of split N applications of 

these two sources vs. the more traditional fall anhydrous N applications. Use of urea and 

UAN solutions may provide greater flexibility in N management than has been 

experienced with anhydrous ammonia. Studies are underway to evaluate the crop and 

water quality effects associated with different N sources and time of application (e.g. see 

reports of work by Gyles Randall and others at the U. of Minnesota: 

http://sroc.cfans.umn.edu/research/soils/index.html). In years when corn growth 

proceeds rapidly, timely side-dressing can be difficult and delayed application can 

severely reduce yields (personal communication, G. Randall, 2007). 

 
182

Application rate. Current N recommendations are usually applied across large 

geographic regions and may provide erroneous results for field-specific soil-crop-climate 

conditions (Gehl et al., 2005; Sawyer and Nafziger, 2005). Grouping soil types with 

similar drainage characteristics, rooting depth, and organic matter content is a feasible 

approach for determining more localized N recommendations and may result in more 

environmentally friendly N management (Oberle and Keeney, 1990). Remote sensing, 

geographic information systems, and variable application technologies offer an 

opportunity to develop and implement site-specific N recommendations, but the 

agronomic understanding of yield response to N on a site- and season-specific basis lags 

behind the technological innovations. There are instances, however, where considerable 

progress has been made in developing site-specific N recommendations (Raun et al., 

2005). 

Application of N near rates that provide the EONR usually results in drainage tile 

flow having nitrate-N concentrations in the range of 10-20 mg/L NO3-N for soybean-corn 

rotations and 15-30 m/L NO3-N for continuous corn (Sawyer and Randall, in press). 

Application of N above the EONR further increases NO3-N losses and reduces net 

economic return. To the extent that N is being applied above the the EONR, reductions 

in N loss through tile drains can be achieved with concurrent positive effects on net 

return (Sawyer and Randall, in press). 

 A review of the effects on N rates on corn-soybean systems in the upper MARB 

was conducted by Sawyer and Randall (in press) who found that in order to achieve a 

30% reduction in tile drainage nitrate-N load, based on a study in Illinois, the N rate had 

to be reduced by 78 kg N/ha (70 lb N/ac) below the EONR, resulting in a large net 

economic loss ($67/ha or >$27/ac). These results illustrate an example of the risk of 

potentially large economic losses to farmers (and their communities) if they are asked to 

reduce N rates below their maximum net return or EONR (Sawyer and Randall, in press). 

The potential environmental benefits of any N rate reductions are highly site-specific, and 

will also depend on how farmer’s past N rates match their site-specific EONRs. 

Economically optimum N rates are not the same across the cornbelt states, and the 

same is true for other crops because of differences among soils, adapted crop varieties, 

climate, management and many other factors that influence production and crop N 

requirements (Hong et al., 2006; Sawyer and Nafziger, 2005). Corn N needs vary widely 

both among and within fields (Scharf et al., 2005; Lory and Scharf, 2003). In some 

fields, in some areas of the MARB, where farmer’s N rates have exceeded the EONR 

(especially where elevated N concentrations have been observed in water resources) there 

may be opportunities to reduce N rates for corn (Mamo et al., 2003) and other crops. 

Nitrogen application rate reductions must be economical for the farmer while also 

protecting water resources. Prior history of many management inputs including fertilizer 

N, manure, and tillage can affect crop N response and EONR interpretations. Farmers 

should carefully consider N rates and evaluate results over several years, in the same 

fields or plot areas. Rate reduction results obtained in one year can be highly affected by 

environmental conditions. For example, it is not uncommon to observe year to year 

variations in rain-fed corn yields ranging above 3.1 to 4.5 Mg/ha (50 to 90 bu/ac), and 
183

economic N rates associated with those yields to vary by more than 60 to 84 kg N/ha/yr 

(54 to 75 lb N/ac/yr) (Sawyer and Randall, in press; Mamo et al., 2003; Jaynes et al., 

2001). 

 As discussed in Section 3.2, higher crop yields (Figure 50) have resulted in 

increased N removal in harvested grain, without increased N fertilization. Greater crop 

harvest N removal may have helped contribute to slight reductions in net N inputs in the 

entire MARB since about 2000; particularly in the Ohio and Upper Mississippi River 

subbasins (see Section 3.2), the two sub-basins which also contribute the greatest annual 

and spring N flux to the NGOM. Increased crop yield trends, improved plant genetic 

selection, and pest control may also be contributing to the reduced nitrate-N transported 

to the NGOM since the mid-1990s, and the steady decline in total N delivered to the 

NGOM since the 1980s (see Section 3.1.1 and Figure 17). Any reductions in N 

application rates could threaten attainment of high crop yields, which are vital to 

profitable production, and which have contributed in some measure to the reductions in 

net N inputs and riverine N discharge mentioned above. 

Figure 50: Average corn yields in six leading corn-producing states (IA, IL, IN, MN, NE, and OH), 1990-

2006 (Source:USDA National Agricultural Statistics Service). 

 Challenges and complexities of determining the EONR in individual fields and 

farms, prevent the ability to make any general conclusions regarding N rate reductions 

across the MARB that will achieve specific N load reductions to the NGOM. Because of 

the complexity and dynamic nature of the N cycle, soil tests for N (nitrate, mineralizable 

N) have not met with much success in practical field applications (e.g. Scharf et al., 

0

2

4

6

8

10

12

19901991199219931994199519961997199819992000200120022003200420052006

Year

Yield, Mg/ha

126 bu/A 

160 bu/A
184

2006a). Some, like the Pre-Sidedress Nitrate Test (PSNT), have resulted in modest 

successes in N rate adjustments, particularly where there is a long history of manure 

applications and there has been a build-up of residual soil N (organic and inorganic). A 

new soil N test (ISNT) developed in Illinois offered promise of more reliably predicting 

mineralizeable soil N pools (Khan et al., 2001; Mulvaney et al., 2001); however, a recent 

report indicates the ISNT does not work well elsewhere (Barker et al., 2006a and 2006 b; 

Laboski et al., 2006). 

 One of the key challenges in managing N in farm fields is to minimize 

unnecessary N applications in low-yielding years and to provide adequate N in highyielding years to meet crop demands. Historically, it has been very difficult for even 

experts to predict residual soil N, recently applied fertilizer N, and mineralized N 

accessible by plants during a given growing season (e.g., Schlegel et al., 2005; 

Shehandeh et al., 2005). Furthermore, the inability to accurately predict the amount, 

intensity, or duration of rainfall in a given year, makes it difficult to adjust N rates each 

year for a specific soil, crop variety/hybrid, tillage system, or cropping system. 

Watershed–scale fertilizer management 

 The first watershed-scale study of changing from fall to spring N application 

involves changes in both rate and timing (Jaynes et al., 2004). The Late Spring Nitrate 

Test (LSNT) is designed to help farmers add appropriate amounts of N in the spring 

instead of fall. Use of the LSNT for corn grown within a 400 ha tile-drained watershed in 

Iowa resulted in at least a 30% reduction of nitrate-N concentrations in tile-drain water. 

The LSNT involved changing timing, rate, and source of N fertilizer. Another Iowa 

study concluded that although watershed-scale implementation of LSNT had the potential 

to reduce nitrate loss through drainage water, it could also increase grower risk, 

especially when above-normal rainfall occurs shortly after the side-dress N is applied and 

N is lost to tile drainage or denitrification (Karlen et al. 2005). Development of 

affordable risk insurance or some other financial incentive by federal, state, or private 

agencies may be needed to stimulate adoption of the LSNT. 

Controlled-release fertilizers 

 Controlled- and slow-release N fertilizers (CRN) are fairly commonly used in 

high-value applications, such as horticultural crop and turf production. Products include 

urea formaldehyde and isobutylidene diurea, and sulfur- and polymer-coated products. 

Use of CRN fertilizer is limited because of the high cost, with world-wide consumption 

less than 1% of all fertilizer N products. However, recent advances have brought some 

CRN products to an economical level for many agricultural crops. Controlled-release N 

fertilizers have the potential to significantly improve N use efficiency, maintain crop 

productivity, and minimize the potential for nitrate loss from fields (Blaylock, 2006). 
185

Effects of N management on soil resource sustainability 

It is well known that soil organic carbon (SOC) storage in Corn Belt Mollisols has 

been decreased by long-term cropping. For instance, in an Iowa study to determine the 

effects of cropping systems on SOC, there was 22 to 49% lower SOC than native prairie 

sampled in fence-rows for all cropping systems that had been in place for 12 to 36 years 

[including continuous corn (CC); corn soybean rotation (CS); corn, corn oats, alfalfa; and 

corn oats alfalfa, alfalfa] (Russell et al., 2005). Current efforts to sequester carbon by 

restoring SOC and to obtain benefits of fertility and tilth associated with higher SOC in 

Mollisols should be considered in achieving nutrient load reductions from these crop 

production systems.

Nutrient management practices need to be assessed for their ability to enhance or 

maintain SOC content in addition to their impact on profit, yield, and water quality 

(Jaynes and Karlen, 2005). A careful review of the literature on this subject is warranted 

because of the potential that fertilizer management to achieve water quality 

improvements may lead to further soil quality degradation. Jaynes and Karlen (2005), 

based on Jaynes et al. (2001), find a partial N mass balance for three fertilizer N levels in 

a corn/soybean rotation on Mollisols in the Des Moines lobe region of Iowa. Tillage 

consisted of either moldboard or chisel plowing in the fall and use of a field cultivator for 

seed-bed preparation and for weed control several times during the early growing season. 

The partial N mass balance shows that the 1X and 2X fertilizer N rates have a negative N 

mass balance and the 3X rate had a positive mass balance. Although the 2X rate (134 kg 

N/ha or 120 lb N/ac on corn, no N applied to soybeans) was the economic optimum, the 

negative N mass balances may indicate a long-term decline in soil fertility. According to 

the authors, “The lower two N rates were thus effectively mining N from the SOM, 

which would result in a measurable decrease in SOM and a degradation of the soil 

resource over the long term." Although all treatments had average nitrate-N 

concentrations above 10 mg/L nitrate-N, there were large and consistent differences 

among N loads in drain tile (Table 15). The 1X and 2X treatments achieved drain tile 

nitrate-N load reductions of 39% and 27%, respectively, compared to the 3X fertilizer N 

rate (201 kg N/ha or 179 lb N/ac). 
186

Table 15: Partial N balance for 4-year rate study by Jaynes et al. (2001). The last two columns added here 

and were not part of original table. 

-------------N inputs------------ ------------------N outputs------------------- 

Fertilizer 

Rate 

Total 

Fertilizer 

Applied 

Total Wet 

and Dry 

Deposition 

Total 

Fixed 

Total 

Grain 

Removed 

Total 

Drainage 

Loss 

Total 

Runoff 

Change 

of 

Residual 

Mineral 

N 

N 

Balance 

Residual 

(Residual/ 

Fixed)*100 (Residual 

/total 

flux)*100 

 -------------------------------------------kg N/ha ----------------------------------------------- -------------%-------------

------ 

1X 144 43 395 522 119 0 6 -55 -14 -4.4 

2X 289 43 397 590 142 0 13 -26 -6.5 -1.8 

3X 414 43 394 606 195 0 -7 47 12 2.8 

 The N mass balance approach to determining long-term changes in SOC or SOM 

presents numerous problems. First, there is no mechanism for lower fertilizer N 

applications to directly stimulate increased SOM mineralization. Any effect on SOC 

would be due to lower residue, particularly during the corn phase of the rotation and 

during soil tillage. Secondly, although a very high quality study, the partial N mass 

balances shown are subject to different interpretations if only small errors exist. For 

instance, the total mass balance residual is less than 5% of the total fluxes measured and 

is 6 to 14% of the estimated N fixation. Therefore, small imprecision in estimated or 

measured values could lead to different interpretations. 

 A number of studies have made direct measurements of SOC over long-term 

studies of fertilizer rates. At least six relevant studies (three in IA and one each in KS, 

MN, NE) have been conducted on Mollisols in the Corn Belt. The general conclusion 

from these studies is that high fertilizer N rates on continuous corn will lead to SOC 

increases and that sub-optimal N rates lead to SOC depletion. There is no direct evidence 

for an effect of lower non-zero fertilizer rates, near the economic optimum, leading to 

decreases in SOC from these studies. 

 Russell et al. (2005) analyzed studies of two Iowa sites (Kanawha and Nashua) 

for the impact on SOC of four N fertilization rates (0, 90, 180, and 270 kg N/ha/yr or 0, 

80, 161, and 241 lb N/ac/yr) and four cropping systems [continuous corn (CC), corn 

soybean (CS); corn-corn-oat-alfalfa (CCOA), and corn-oat-alfalfa-alfalfa (COAA)]. One 

study had been ongoing for 23 years and the other for 48 years at the time of sampling of 

SOC in 2002. The only difference related to fertilizer rate was for the 23 year experiment 

(the Nashua site). In this experiment, the 270 kg N/ha/yr (241 lb N/ac/yr) for CC had 

higher SOC for only the 0-15 cm (0-5.9 in) depth. There were no differences among the 

0, 90, and 180 kg N/ha/yr rates for CC at the Nashua site for any depths. There were also 

no differences for the 0-100 cm (0 – 39 in) soil for any N rates used for CC, including the 

highest rate of 270 kg N/ha/yr (241 lb N/ac/yr). There were no other significant fertilizer 

N rate effects found in the study (Russell et al., 2005). 
187

 An earlier Iowa study that included the Nashua and Kanawha sites and a third site 

(Sutherland) reached similar conclusions as those of Russell et al. (2005). In that study, 

Robinson et al. (1996) found that N fertilizer rate on corn (0-180 kg N/ha/yr or 0 -161 lb 

N/ac/yr) was not significant in determining SOC but only whether fertilization occurred. 

In both studies (Russell et al., 2005; Robinson et al., 1996), the cropping systems with 

alfalfa [termed meadow in Robinson et al. (1996)] had the highest SOC. Corn silage 

treatments and no fertilizer treatments had the lowest SOC (Robinson et al., 1996). A 

third Iowa study did not compare SOC under different fertilizer rates but did show that 

high fertilizer N (206 kg N/ha/yr or 184 lb N/ac/yr) resulted in increases in SOC over 15 

years with continuous corn (Karlen et al. 1998a). The general conclusion from the Iowa 

studies is that for either CC or CS systems, fertilizer rate has little or no effect in the 90-

180 kg N/ha/yr (80-161 lb N/ac/yr) range. Given that the average N fertilizer application 

to corn in Iowa was 158 kg N /ha (141 lb N/ac) in 2005 (USDA ERS: 

http://www.ers.usda.gov/Data/ARMS/app/CropResponse.aspx) and the economic 

optimum rate ranged between 67 and 172 kg N/ha or about 60 and 154 lb N/ac 

(approximate mean of 137 kg N/ha or 122 lb N/ac) during 1996 and 1998 in the Iowa 

study by Jaynes et al. (2001), it seems unlikely that these rates would lead to a depletion 

of SOC due to a N rate effect. Corn yields with the moderate N rates in the Jaynes et al. 

(2001) study ranged around 10 Mg/ha (159 bu/ac) and the Iowa state average corn yield 

in 2005 was about 10.9 Mg/ha (173 bu/ac). 

 Results from other studies in the Corn Belt are mixed and have found no 

consistent effect of N rate on SOC. In Kansas, Omay et al. (1997) found no effect of 

either 224 or 252 kg N/ha (200 or 225 lb N/ac) versus no N for over 10 years of CC or 

CS. A small significant difference in SON (less than 5% decrease) was found on one soil 

for the 0 N treatment. Increased residue inputs were attributed to N fertilization, and 

inclusion of soybean in the rotation reduced SOC and soil organic N. In contrast, CC 

receiving 200 kg N/ha/yr (179 lb N/ac/yr) for 13 years had higher SOC than in the 0 N 

treatment on a Minnesota Mollisol (Clapp et al., 2000). In an 18-year experiment in 

Nebraska, N rate (0, 90, 180 kg N/ha/yr or 0, 80, 161 lb N/ac/yr) had an effect on SOC in 

the 0 to 7.5 cm (0 to 2.9 in) soil after eight years but had no effect after 18 years, 

presumably due to tillage differences (Varvel 2006). 

 Recent work in Nebraska on an irrigated Mollisol compared long-term (initiated 

in 1999) continuous corn and corn-soybean rotations under recommended and intensive 

management and found that SOC was increased under recommended and intensive 

management of CC but not in the CS systems (Adviento-Borbe et al., 2007; Dobermann 

et al., 2007). These scientists also reported that greenhouse gas (GHG) emissions from 

agricultural systems can be kept low when management is optimized towards better 

exploitation of the yield potential. To accomplish SOC increases while keeping GHG 

emissions low, Dobermann et al. (2007) reported the following required factors: 1) 

choosing the right combination of adapted varieties or hybrids, planting date, and plant 

population to maximize crop biomass production; 2) tactical water and N management, 

including frequent N applications to achieve high N use efficiency and minimized N2O

emissions; and 3) a deep tillage (non-inverting) and residue management approach that 

favors a build-up of SOC as a result of large amounts of crop residues returned to the soil. 
188

 If a fertilizer effect on SOC exists, it is more likely to occur under CC than CS 

because increased fertilizer generally leads to increased corn production. It is logical to 

assume that increased corn production (including grain, stover, and roots) should lead to 

increased SOC. In general in the published studies, this relationship does not hold, 

although applying zero N fertilizer generally leads to less SOC over time than high 

fertilizer N rates. In summary, although it is beyond the scope of the SAB Panel to 

review all the research relevant to changing SOC in Corn Belt soils, it is clear that 

inclusion of alfalfa in a rotation is very effective at building SOC. The effects of tillage 

are not clear. Based on the existing literature, there is evidence that changes in fertilizer 

rates within the range of those optimum for corn production are unlikely to lead to longterm SOC and SON declines. Although it is possible to build SOC under CC with 

relatively high fertilizer additions [e.g., 201 to 299 kg N/ha/yr or 179 to 267 lb N/ac/yr 

(Adviento-Borbe et al., 2007; Dobermann et al., 2007) and 206 kg N/ha/yr ( 184 lb 

N/ac/yr) Karlen et al. 1998b) care must be taken to ensure that these fertilizer additions 

are sustainable economically and that they do not harm water quality. From a global C 

balance perspective, it is also worth noting that there is a C emissions cost of producing 

N fertilizer that would need to be taken into account when doing C mass balances for 

higher fertilizer N rates on corn. However, if high-yield production is achieved, with 

good N use efficiency, these fertilizer C emissions may be offset (Adviento-Borbe et al., 

2007). More research on the net effects of N fertilizer rates on SOC and GHG emissions 

is needed. 

Precision agriculture management tools for Nitrogen 

 Global positioning system (GPS) and geographic information system (GIS) 

technologies are becoming more widely adopted by farmers and show promise for 

developing management zones in fields that could target application rates for low- versus 

high-yielding areas (Schlegel et al., 2005) and reduce N applications in areas of the field 

most prone to N losses (Chua et al., 2003). Field-transect apparent electrical conductivity 

(ECa) or electromagnetic induction measurements can help define management zones, 

based on surrogate detection of soil texture differences (Davis et al., 1997; Kitchen et al., 

1999). Reductions in N application rates for corn range from 6 - 46% when using site 

specific management zone approaches as opposed to a uniform rate of N application 

(Koch et al., 2004). Dividing fields into a few management zones might reduce N loss, 

but because of within-field variability, more spatially intensive N management might 

provide greater economic and environmental benefits (Hong et al., 2006; Scharf et al., 

2005). 

 Basing N applications on past yields has not proven to be an effective approach to 

variable rate fertilization of N (Murdock et al., 2002; Scharf et al., 2006b). In-season 

crop N sensing research (chlorophyll meter, remotely-sensed multispectral color images, 

on-the-go and hand-held optical reflectance sensors) (Scharf et al., 2006a), using 

reference “N- rich” or calibration strips or plots in targeted areas within fields (Raun et 

al., 2005) has shown the potential benefits of these newer technologies in providing inseason guidance to farmers and crop advisers for improved N nutrition management. 
189

This “N-rich” calibration approach appears to have been more successful with winter 

wheat than for corn, to date. Chlorophyll meters and remotely sensed crop reflectance 

have been used as an index for plant N status, and N-fertilizer use efficiency improved 

when these techniques were used (Osborn et al., 2002; Varvel et al., 1997). Crop Nsensing technologies present opportunities to reduce and better time fertilizer N 

applications; however, there have been few direct assessments of impacts of these 

approaches on residual soil N and nitrate losses. Further verification of the performance 

of these techniques is needed in order for implementation by farmers to be more 

widespread. 

 When technology costs are considered, economic returns to farmers are often 

inadequate to justify adoption of variable rate N management. Frequently, the costs of 

spatial N management technologies exceed the cost of the fertilizer N saved, which are 

dependent on fertilizer prices. As a consequence, adoption of these technologies has 

proceeded at a slower rate than anticipated, partly because of high technology and 

equipment costs and spatially variable economic returns. Economics research suggests a 

number of reasons for this low slow adoption including high fixed costs of adoption and 

uncertainty in returns. These factors suggest that incentives to encourage adoption may 

need to cover option values and that revenue insurance programs to address the risk may 

be appropriate instruments (Khanna et al., 2000; Khanna, 2001; Isik and Khanna, 2002, 

and Isik and Khanna, 2003). 

 Incentives have been used in Missouri in cost-sharing some of the expenses of 

precision technologies within the USDA EQIP program (Agronomy Technical Note MO35, September 2006). Cost-share in this Missouri USDA NRCS Code 590 nutrient 

management program provides a farmer $49/ha ($20/ac) per year for a three-year 

contract, with the full $148/ha ($60/ac) provided at the end of the first year. Farmers in 

this Missouri EQIP precision N-sensing program are advised to follow guidance for Nsensing interpretation based on work by Scharf et al. (2006a and 2006b). 

Precision Agriculture Management Tools for Phosphorus 

 Spatial variability in soil test phosphorus (P) levels can be large, with levels often 

ranging from very low to very high (agronomic interpretation) in the same field 

(Bermudez and Mallarino, 2007; Wittry and Mallarino, 2004; Reetz et al., 2001; McGraw 

and Hemb, 1995). This variability can also be large in fertilized, manured, and grazed 

pastures (Snyder and Leep, 2007; Mallarino and Schepers, 2005). With the advent of 

commercially available GIS and GPS technologies in the early 1990s, crop advisers and 

farmers began to more precisely define the spatial variability of soil fertility levels, 

including soil test P (Figure 51). In recent years, zone or grid (e.g., 0.25 to 1 ha or .6 to 

2.5 ac) sampling has been used to better define management units to receive different P 

application rates (Reetz et al., 2001), as opposed to the formerly recommended practice 

of whole-field composite sampling (e.g., Thom and Sabbe, 1994). In spite of 

considerable research effort, no widely accepted standard for soil sampling fields for 

precision or site-specific management has been established (Mallarino and Schepers, 

2005), because soils are naturally heterogeneous and their spatial variability occurs at 
190

many scales. Recent soil sampling summary results for more than 3.3 million soil 

samples in North America from both public and private soil testing laboratories also 

showed wide variability in soil test P levels within and among states in the U.S. 

(PPI/PPIC/FAR, 2005). Snyder (2006) summarized the soil test results for the 20 major 

MARB states (over 2.1 million samples) and reported 1) 40% of the states have 

experienced a decline in soil test P since 2001, and 2) 78% of the samples tested below 

50 mg/kg (ppm) Bray 1 equivalent-extractable P and 94% tested 100 ppm or below. In 

fact, crop harvest removal of P exceeds fertilizer plus recoverable manure P in 11 of the 

20 states (PPI/PPIC/FAR, 2002). These data are in agreement with the trends in net 

anthropic P input in the MARB, discussed in Section 3.2 of this report. 

0 10 20 30 40 50 60 70 80 90

Range 1 - 93 ppm

Avg. 28

0 10 20 30 40 50 60

Range 1 - 59 ppm

Avg. 15

Bray 1 Soil Test P, 0 - 6 inches

Fall Soil Samples 2001

W. A. Field R.W. Field

Figure 51: Variability in soil test P levels in typical farmer fields in Minnesota (2007 personal 

communication with Dr. Gary Malzer, University of Minnesota) 

Early season detection of corn P deficiency may be possible with remote sensing, 

but detection of deficiencies later in the season, which correlate better with crop yield, 

has not been successful (Osborne et al., 2002). At this time, remote sensing or on-the-go 

sensing of plant P status does not appear to be as commercially viable as plant N sensing. 

 Variable-rate fertilization can result in better P fertilizer management. For 

example, Burmedez and Mallarino (2007) found that variable-rate technology applied 12 

to 41% less fertilizer and reduced soil test variability on farmer’s fields in Iowa, 
191

compared with the traditional uniform rate fertilization method. Perhaps one of the most 

important aspects of intensive soil sampling and variable-rate P application technologies 

is the capability to apply P fertilizer where it is needed while minimizing or reducing P 

applications in field areas which have elevated soil test P. In Iowa, variable-rate P 

application helped decrease soil test P in field areas with high soil test P, when applying 

manure (Figure 52) or fertilizer (Figure 53). As of yet, however, variable rate or 

precision P fertilization has been shown to have little economic benefit in the major corn 

and soybean producing states compared to uniform applications (Lambert et al., 2006; 

Mallarino and Schepers, 2005). Further, there are ongoing efforts to update soil test P 

crop response calibrations and fertilizer recommendations to optimize P fertilization 

(Beegle, 2005). Soil-Test P Change (mg kg-1)

-3

-2

-1

0

1

2

3

4

5

6

7

8

9

VERY LOW

SOIL P

OPTIMUM

SOIL P

LOW

SOIL P

> OPTIMUM

SOIL P

Variable rate

Uniform rate

Manuring Method

No manure

Figure 52: Effect of variable-rate versus uniform rate application of liquid swine manure on changes in soil 

test phosphorus in Iowa fields [2007 personal communication with Dr. Antonio Mallarino, Iowa State 

University and Wittry and Mallarino (2002)]. 
192

Figure 53: Effect of variable rate versus uniform rate application of fertilizer P on soil test P in multiple 

Iowa fields across multiple years (2007 personal communication with A. Mallarino, Iowa State University). 

 Numerous studies have shown a strong relationship between soil test P levels and 

the concentration of dissolved P in runoff (Sharpley et al., 2006a, 2006b; Andraski and 

Bundy, 2003; Pote et al., 1999) and tile drainage (Heckrath et al., 1995). Recent work by 

Gentry et al. (2007) showed that tile drainage P losses in Illinois can exceed one kg 

P/ha/yr (0.9 lb P/ac/yr), with much of the loss occurring during a few peak storm events 

in the spring. However, annual manure or fertilizer P applications can control the 

concentration of total and dissolved P in surface runoff (Pierson et al., 2001; Sharpley et 

al., 2001). 

 Soil test P thresholds alone cannot define the potential or risk of P losses from 

agricultural fields. Slope, hydrologic characteristics, tillage, P rate, and time after P 

application before a runoff producing rainfall, and other factors also affect the risk of P 

loss (Sharpley et al., 2006a). Soil test P thresholds alone cannot define the potential or 

risk of P losses from agricultural fields. Slope, hydrologic characteristics, tillage, P rate, 

and time after P application before a runoff producing rainfall, and other factors also 

affect the risk of P loss (Sharpley et al., 2006a). To address all factors influencing P loss 

from agricultural fields, an environmental risk assessment tool (the P Index) was 

proposed by Lemunyon and Gilbert (1993), which has been regionally modified and 

adopted by 49 of 50 states in the U.S. to identify and delineate the risk for agricultural P 
193

loss for use in the development of Comprehensive Nutrient Management Plans (Sharpley 

et al., 2003). Use of P Indices has also been encouraged by industry, in recognition of the 

spatial variability in soil test P levels within fields, and the spatial variation in source and 

transport factors (Snyder et al., 1999). “Variable rate P application can be practically 

implemented on the basis of P index ratings for field zones, not just based on soil test P” 

(Wortmann et al., 2005). Variable rate fertilizer P application is becoming more common 

in Nebraska, Iowa, Missouri, Kansas, and other states, and some custom applicators are 

beginning to apply manure at variable rates. 

Nutrient management planning strategies 

 A survey of 127 farms (90% of all farms) in two northeastern Wisconsin 

watersheds offers some insight into how successful nutrient management has been in 

reducing nutrient applications (Shepard, 2005). Farmers with a nutrient management 

plan (53% of farms) applied less N and P (139 kg N/ha and 31 kg P/ha or 124 lb N/ac and 

28 lb P/ac) than farms without a plan (188 kg N/ha and 44 kg P/ha or 168 lb N/ac and 39 

lb P/ac), but only half the farmers credited on-farm manure N, and only 75% fully 

implemented their plans on most of their acres. 

 For nutrient management planning to decrease nutrient loss, technical and 

financial assistance programs need to focus on plan implementation and maintenance in 

the MARB rather than on targeting the number of plans written in a given period. 

Despite programs subsidizing plan writing, a critical limitation is the lack of certified 

plan writers to meet the demand and deadlines. Further, there needs to be an effective 

mechanism to ensure plan adoption and regular updating of plans. Efforts are underway 

in the Heartland states of the MARB (IA, KS, MO, NE) to develop nutrient management 

plan assessment protocols. This aims to identify key factors that limit plan 

implementation so that practical solutions can be developed. One option is preparation of 

a simplified plan that farmers can quickly refer to. Also, documenting nutrient 

management plan implementation is being rewarded with financial credits in New York 

drinking water supply watersheds (Watershed Agricultural Council, 2004). These credits 

can be used to purchase or upgrade equipment that would need to be used to implement 

the plan, such as manure spreaders, injectors, etc. 

 An assessment is needed of the socioeconomic barriers to successful adoption of 

nutrient management planning strategies in the MARB as well as the N and P loss 

reductions achievable. Such an assessment has been done in a drinking water supply 

watershed for New York City that claims a 93% participation in volunteer conservation 

programs (Watershed Agricultural Council, 2004). A survey of CREP participants 

showed they were generally older and more likely to obtain information from extension 

agents, consultants, and watershed council personnel than non-participants, but there was 

no difference in educational level or farming status (full or part time) (James, 2005). 

Overall, negative attitudes toward voluntary adoption of BMPs were a result of the loss 

of productive land and loss of being able to decide independently what to do on their own 

land. These survey results illustrate the difficulties in gaining adoption of nutrient 
194

management BMPs by farmers in any watershed, transferring new BMP technology, and 

the socioeconomic pressures faced. 

Key Findings and Recommendations 

 Reductions in N losses and residual soil NO3-N are possible with attention to 

improved in-field N management. It may be possible to reduce N rates and alter N 

timing in some portions of the MARB. Such rate reductions may be accomplished 

through implementation of refined management, but they must be economical for farmers 

and care must be taken to protect soil resource sustainability. Crop N sensing and 

variable rate N management implementation, using management zone approaches may 

prove useful in attainment of economic optimum N rates in individual fields, which may 

also help reduce N losses. Higher fertilizer, fuel, and machinery costs have stimulated 

increased interests in some newer N management technologies, as well as other means to 

improve fertilizer N effectiveness and efficiency; however use of site-specific or 

precision technologies has not yet proven financially rewarding to many farmers, due to 

the high cost of sampling, ground- truthing, and application technology. Based on these 

findings, the SAB Panel offers the following recommendations. 

• Because of the importance of both N and P to Gulf hypoxia and as various 

cropping systems can have different positive and negative effects on N and P 

export reduction, remedial strategies must be directed at system-wide nutrient 

management rather than either N or P applications alone. Future research to 

evaluate the effects of different nutrient management impacts on crop production 

should include measures of water and air quality effects. 

• There is a lack of consistent year-to-year USDA nutrient management survey 

data, which hinders any broad nutrient use and management evaluation and 

interpretations. These data will become more important in monitoring and 

understanding changes in nutrient management practices as biofuel markets 

expand. Consistent year-to-year data collection on nutrient management of major 

crops and emerging energy crops is recommended. 

• Cost-share incentives like the USDA payment support for crop N –sensing and 

precision N management in Missouri, intensive educational programs (e.g., onfarm demonstrations), and/or other means should be explored to encourage the 

agricultural community to improve nutrient use efficiency and effectiveness with 

all nutrient sources (i.e., fertilizer, manure, biosolids, composts, by-products, 

etc.). Such programs may be especially helpful in corn systems in the upper 

Mississippi and Ohio River subbasins, which have been identified as major 

contributors of spring nitrate-N flux to the NGOM. 

• Although the economic and water quality impacts of controlled release fertilizers 

in commercial field crop systems have not been fully proven, their beneficial use 

should be explored through additional research and demonstrations at field and 

watershed scales. Programs to stimulate greater adoption of locally-proven 
195

technologies like urease and nitrification inhibitors (and controlled release 

fertilizers, once proven economically and environmentally effective) to enhance 

crop nitrogen recovery and use efficiency, should be considered as the shift 

towards greater urea and urea-ammonium nitrate N use continues. 

• Watershed-scale evaluations of split applications of N in the spring for corn 

should be conducted to determine watershed-scale benefits of this N management 

approach compared to the more traditional application of anhydrous ammonia in 

the fall, especially in the upper Mississippi and Ohio River subbasins. 

• More research on the net effects of N fertilizer rates on soil organic carbon (SOC) 

and greenhouse gas (GHG) emissions is needed. 

• Crop and animal production systems are essential to the economic viability of 

agriculture in the MARB. Thus, an infrastructural assessment of how animal 

production can co-exist with grain and forage production is needed. Long-term 

strategies should be explored whereby more effective crop and animal production 

systems remedy or avoid excessive N and P loading to water and air resources. 

• Cost-benefit ratios vary among farmers; with for example, labor availability, farm 

organization, and financial situation. However, past experience shows that 

adoption of conservation practices is not solely dependent on cost-effectiveness. 

Thus, there needs to be consideration of the socioeconomic barriers to, and 

impacts of, adoptions of nutrient management planning strategies in the MARB. 

New approaches should be investigated to overcome socioeconomic barriers, 

including incentive programs. 

4.5.7. Effective Actions for Other Non-Point Sources 

Atmospheric Deposition 

 This section reviews actions for reducing NOx emissions that contribute to 

atmospheric deposition of nitrogen. For the United States as a whole, atmospheric 

deposition of oxidized nitrogen compounds released during fossil fuel combustion 

contributes an estimated 30% of the entire inputs of new nitrogen (Howarth et al., 2002). 

As discussed earlier in Section 3.2, atmosphere deposition of oxidized nitrogen is less 

important in the MARB but still accounts for an estimated 8% of nitrogen contributions 

to the upper MARB and 16% of the nitrogen inputs to the Ohio River basin. NOx 

emissions to the atmosphere in the United States could be virtually eliminated at 

reasonable cost using currently available technologies (Moomaw, 2002; Howarth et al., 

2005). In addition to potential benefits concerning Gulf hypoxia, reducing NOx 

emissions in the MARB can contribute to improved local air and water quality and can 

reduce atmospheric transport of nitrogen to the northeastern States, where atmospheric 

deposition is an even more significant problem. 
196

In addition to deposition of oxidized nitrogen, there is significant deposition of 

ammonia and ammonium (NHx) in some regions of the MARB. These are not 

considered in the mass balance approach for nitrogen in Section 3.2 because the NHx 

originates largely from volatilization from animal wastes and other agricultural sources 

and so does not represent new nitrogen inputs to the basin, but rather a recycling of 

nitrogen within the basin (Howarth et al, 1996). Nonetheless, high rates of volatilization 

followed by conversion to ammonium nitrate or sulfate can lead to significant longdistance transport and contribute to reactive N distribution in other sensitive areas. 

Furthermore, high rates of NHx deposition in the basin can result in increased leakage of 

nitrogen to downstream aquatic ecosystems. In Iowa, Minnesota, and Wisconsin, NHx 

deposition exceeds NOy deposition, and averages over 7.5 kg N/ha/yr (6.7 lb N/ac/yr) in 

Iowa (results from CMAQ model, Robin Dennis, NOAA, unpubl.) 

Mobile sources account for approximately 55% of NOx emissions to the 

atmosphere on a national level (Melillo and Cowling, 2002). While automobiles have 

been subject to fairly strict NOx standards in recent years, emissions from light trucks 

have not historically been as strict. Tightening regulations on light trucks represents an 

opportunity for significant reduction in NOx emissions, as approximately half of new 

vehicle sales in recent years have been light duty trucks (Moomaw, 2002). Heavy diesel 

trucks, buses, and trains have accounted for a growing fraction of NOx emissions because 

of strict NOx standards on automobiles and the absence of similarly strict controls on 

heavy diesel vehicles. 

Stationary sources account for approximately 45% of NOx emissions, with 

electric generating facilities accounting for roughly half of all stationary source 

emissions, and industrial fuel combustion account for slightly less than one-third. The 

remainder of stationary-source NOx emissions are from non-fuel industrial processes 

(12%) and from commercial, institutional, or residential fuel combustion (8%) (U.S. 

EPA, 2006a). 

 Stringent new source performance standards have greatly reduced emissions from 

new electric generating facilities. Low emission, combined-cycle gas turbines account 

for most new electric generating capacity in recent years (Bradley and Jones, 2002). 

Unfortunately, some existing policies provide incentives that discourage more 

widespread adoption of new, cleaner technologies. For example, under the Clean Air 

Act, high NOx emissions by older, coal fired power plants are “grandfathered,” and 

therefore not subject to the stringent emission standards of new generating capacity. As a 

consequence, electric utilities have the incentive to keep older coal plants running far 

beyond what would otherwise be their economic lifespan (e.g., Ackerman et al., 1999; 

Nelson et al., 1993; Maloney and Brady, 1988). As a result, while 90% of new electric 

generating capacity is produced with gas turbines, coal still produces 55% of the 

electricity in the US (Moomaw, 2002). And it was estimated that in 1998, coal-fired 

power plants were responsible for nearly 90% NOx emissions from electric power 

generation (U.S. EPA, 2000b; U.S. EPA, 2006a). About a quarter of the coal-fired 
197

electric generating capacity in 1996 was constructed prior to 1965, and almost one-half 

was constructed prior to 1975 (Ackerman et al, 1999). 

Considerable reductions in NOx emissions can be achieved with existing 

commercial technologies by replacing outdated coal-fired capacity with modern gas-fired 

combined-cycle power plants (Howarth et al., 2005). Existing coal plants can also be 

retrofitted with new control technologies, such as Low-NOx burners (Bradley and Jones, 

2002; Ackerman et al., 1999). Other promising technologies for reduction emissions 

from coal-fired power plants include fluidized bed boilers (Co-Generation Technologies, 

2006), and gasified coal combined-cycle power plants, and sequestration of emissions 

(U.S. DOE, 2006). 

 For the most part, NOx emissions in the United States are regulated because of 

concerns over formation of smog and ozone and seldom because of water-quality 

concerns (Moomaw, 2002; Melillo and Cowling, 2002). Since smog and ozone pollution 

occur mostly in summer months, regulation of NOx emissions from stationary sources 

has often focused on summer-time only regulation (Howarth et al. 2005). Since the 

largest cost of controlling NOx from power plants is the capital cost of building scubber 

systems, the additional cost of requiring year-round NOx control from power plants is 

small compared to that for summer-time only controls. Thus, year-round operation of 

existing control technologies represents a cost effective approach for reducing NOx 

emissions. Some local and state governments, such as New York State, have recently 

moved towards year-round regulation of NOx because of concern over coastal nitrogen 

pollution (Ron Entringer, NY State DEC, personnel communication). 

Residential and Urban Sources 

 Urban and suburban runoff comes from a variety of sources, including impervious 

surfaces like roads, rooftops and parking lots, as well as pervious surfaces like lawns. 

Urban and suburban runoff can be important sources of pollutants, especially for local 

water quality effects. For example, the National Water Quality Inventory: 2000 Report 

to Congress concluded that urban runoff is a major source of water quality impairment in 

surface waters (U.S. EPA, 2002). There are a variety of actions to control non-point 

urban sources, including both structural and non-structural practices (e.g., U.S. EPA, 

2005). 

 Although controlling urban non-point sources can provide significant benefits 

from improvements to local water quality, these non-point sources are not significant 

determinants of hypoxia in the Gulf of Mexico, both because concentrations tend to be 

lower than those from agricultural sources and because the urban land comprises less 

than 1% of the Mississippi River basin (e.g., Mitsch et al., 1999). Thus, although actions 

to reduce urban non-point sources may be justified, these control actions will not likely 

contribute significantly to reductions in the size of the Gulf of Mexico hypoxic zone. 

Since control of urban non-point sources will not have an important role in reducing 

hypoxia, we do not focus on actions to reduce urban non-point sources of nutrients in this 

report. 
198

Key Findings and Recommendations 

 Atmospheric deposition is a small but significant (8% in Upper Mississippi and 

16% in Ohio River subbasins) contribution to N inputs in the Mississippi River basin. 

Opportunities exist to lower NOx emissions in a number of ways, but it is not likely that 

hypoxia will drive most of these regulatory decisions. Rather, hypoxia reduction and 

other water quality benefits should be incorporated in a number of regulatory decisions 

regarding air pollution. Based on these findings, the SAB Panel offers the following 

recommendation. 

• Water quality benefits and effects on hypoxia should be incorporated into 

decisions involving retirement or retrofitting of old coal-fired power 

plants, NOx controls such as the extension of the current summertime 

NOx standards to a year-round requirement, and emissions standards and 

mileage requirements for sport utility vehicles, heavy trucks and buses. 

4.5.8. Most Effective Actions for Industrial and Municipal Sources 

Sewage treatment plants and industrial dischargers represent a more significant 

source of N and P in the MARB than was originally identified in the Integrated 

Assessment. Although most point sources in the MARB do not have permits that require 

removal of N or P from discharged effluent, as local water quality standards for these 

nutrients have not yet been developed, states are charged with developing water quality 

criteria for achieving and maintaining designated beneficial uses of surface waters, 

including those waters that receive sewage treatment plant effluent. However, the 

process by which these criteria are translated into quantitative and enforceable nutrient 

limits from regulated point sources remains unclear. 

Based on data from the recent MART (2006b) report, the SAB Panel has 

estimated that permitted point-source discharges represented approximately 22 and 34% 

of the average annual total N and total P flux to the Gulf, respectively, for the 2001 to 

2005 water years (for a detailed discussion see Appendix D: Calculation of Point Source 

Inputs of N and P). These point sources represent a significant opportunity to reduce N 

and P loadings that should be fully evaluated in the context of other potential 

management changes in the MARB. 

Encouraging behavioral changes of non-domestic sewer users as well as 

increasing capital investments in sewage treatment and industrial treatment plant 

upgrades have proven to be effective approaches to managing nutrient discharges in other 

areas of the U.S. (U.S. EPA, 2004b; U.S. EPA, 2003a; Chesapeake Bay Commission, 

2004). The use of Biological Nutrient Removal and Enhanced Nutrient Removal 

technologies for N and P removal are being implemented to reduce N and P 
199

concentrations in sewage treatment plant effluent discharge by 50 to 80% (Maryland 

Department of Environment, 2005; U.S. EPA, 2004b). Sewage treatment plant upgrades 

designed to remove phosphorus typically include enhanced chemical precipitation 

applied alone or in combination with biological phosphorus treatment and membrane 

filtration. These types of sewage treatment plant unit operations, which can achieve 

effluent discharge phosphorus concentrations as low as 0.1 mg/L total phosphorus or less, 

now constitute the BMP for phosphorus removal at sewage treatment plants. Removing P 

to a 0.1 mg P/L limit is most commonly implemented where there is a market for water 

recycling, such as in communities located in the desert Southwest, and the increased cost 

can be justified. In locations where there is no market for recycled water, higher limits for 

P (for example, 0.3 or < 1.0 mg P/L) will be more cost effective. 

 The SAB Panel presents an example calculation to demonstrate the magnitude of 

reduction possible in riverine total N and P fluxes to the NGOM if technology for N and 

P removal from sewage effluent were implemented for large sewage treatment plants (0.5 

million gallons per day and above) across the MARB. Based on the SAB Panel’s 

adjustment to the MART report’s estimates of N and P effluent from sewage treatment 

plants (MART, 2006b), the SAB Panel has calculated that upgrades for large sewage 

treatment plants in the MARB to achieve total N concentration limits of 3 mg/L could 

create reductions in N flux from sewage treatment plants from 192,000 metric tonne N/yr 

(212,000 ton N/yr) to 70,000 metric tonne N/year (77,000 ton N/yr), about a 64% 

reduction in annual N flux from sewage treatment plants. This translates into a reduction 

of total annual N flux to the Gulf by about 10% and the total spring N flux by about 6%. 

Upgrading to achieve P concentrations of 0.3 mg/L would create reductions in P fluxes 

from sewage treatment plants from 41,000 metric tonne P/yr (45,000 ton P/yr) to 10,500 

metric tonne P/yr (11,600 ton P/yr) or about a 75% reduction in annual flux from sewage 

treatment plants to the MARB. These reductions, in turn, would translate into a decrease 

in the total annual P flux to the Gulf by about 20% and the total spring P flux by about 

15%. It is important to recognize that these estimates assume that the changes in 

biosolids quality and production rates resulting from the capital improvements to the 

sewage treatment plant do not adversely impact nutrient management procedures 

implemented at biosolids land application sites. 

In the Chesapeake Bay watershed, nutrient reductions from sewage treatment 

plant upgrades were determined to be as cost effective as, and more predictable than, the 

estimated reductions achieved through implementation of agricultural non-point source 

BMPs. The Chesapeake Bay Commission (2004) found average point source costs to 

remove N and P of $8.56/lb and $75/lb, respectively, which was within the range of most 

widely implemented agricultural BMPs (U.S. EPA, 2003b). The Commission stated that 

“this technology-based approach provides the highest degree of confidence for consistent, 

long-term reductions. Furthermore, the cost of this technology has continued to decline 

in recent years.” 

However, there are many differences in point source distribution, population, and 

income in various sub basins of the MARB compared to other areas of the country where 

point sources have had total N and P reductions (such as the Chesapeake Bay or Long 
200

Island Sound). Therefore, a cost effectiveness analysis of point-source controls of N and 

P in the MARB is needed to fully evaluate this particular method of reducing nutrient 

inputs to rivers in the context of non-point source control costs. A part of that analysis 

should consider the cost of N and P removal that could be optimized by establishing 

loading caps for individual treatment plants and/or groups of plants within river basins 

and by allowing nutrient credit trades between the plants. This “point-to-point” trading 

allows those plants that can most efficiently achieve reductions to sell nutrient reduction 

credits to plants that would incur much higher costs to achieve their loading cap. This 

approach is being used in Long Island Sound and in the Chesapeake Bay watershed 

within Virginia. These point-to-point trading programs are consistent with an overall cap 

and trade program as discussed in Section 4.4.3. 

 Another potential approach for reducing the nutrient discharge from sewage 

treatment plants, that could be applied alone or in combination with plant upgrades, is to 

encourage local sewer districts to establish more stringent nutrient pretreatment standards 

for private industries and other non-domestic sewer users. Meat packing, chemical 

manufacturing and food processing are examples of the types of industries that generate 

wastewater containing large amounts of N and P. Through the regulatory authority 

granted to them under the National Pollutant Discharge Elimination System (NPDES) 

program, sewer districts can encourage industries to reduce their nutrient discharge to 

sewage treatment plants through the establishment of local sewer discharge nutrient 

limits as well as by the judicious development of technology-based wastewater surcharge 

rates. 

The overall decrease in the mass of nutrients discharged into the local sewer 

system due to pretreatment will improve the quality of both the sewage treatment plant 

effluent and biosolids and will result in a net reduction of nutrients entering the MARB. 

A feasibility study is needed to evaluate the regulatory and economic options that could 

be applied to provide incentives for major industries to identify and implement pollution 

prevention measures to reduce and/or recycle nutrients that would otherwise be 

discharged into the local sewer system. 

 In addition, industrial treatment plant upgrades designed to remove nutrients can 

also reduce nutrients that are directly discharged to the MARB and the Gulf. Industrial 

discharges account for about 28% of the point source N flux and 23% of the point source 

P fluxes, or 75,000 metric tonne N/yr (83,000 ton N/yr) and 17,000 metric tonne P/yr 

(18,700 ton P/yr). Experience in other regions has shown that industrial sources could be 

targeted on a permit-by-permit basis since frequently a limited number of permitted 

facilities are responsible for a large part of the load. This approach could be 

recommended for the MARB. It would be useful to design initial efforts to focus on 

discharge categories likely to have high nutrient discharges. Examination of discharge 

information (Table 3, MART, 2006b) reveals that two categories (industrial organic 

chemicals and plastic materials/synthetic resins) account for about half of industrial N 

discharges, about 45,000 metric tonne N/yr (50,000 ton N/yr). For P, four categories 

(crude petroleum and natural gas, electrical services, refuse systems, and wet corn 

milling) account for about 40% of the industrial load or about 5,500 metric tonne P/yr 
201

(6,000 ton P/yr). Industries in these categories should be evaluated for opportunities to 

reduce N and P discharges through pollution prevention, process modification or 

treatment. 

While P removal is technologically feasible and widely implemented elsewhere, 

advanced treatment increases the amount of biosolids generated and, therefore, the land 

area needed to manage a given amount of biosolids based on P and N needs of the crop, 

rather than just the N requirements. This will create additional costs for biosolidsmanagement programs in the MARB and needs to be considered when evaluating the 

total cost of implementing P removal at sewage and industrial treatment plants in the 

basin. 

Unlike nitrogen, which can be biochemically transformed and removed from the 

sewage treatment plant as a volatile gas (N2 and/or N2O) through the 

nitrification/denitrification process, phosphorus is simply moved from the liquid to solid 

phases and accumulates in the biosolids. Physical upgrades in sewage treatment plants 

specifically aimed at reducing the phosphorus concentration in the effluent discharge 

typically include substantial additions of precipitating chemicals (e.g., alum) alone, or in 

combination with, higher efficiency membrane filtration. The net effect of these capital 

improvements is a significant increase in the mass of biosolids requiring handling and 

management. Most biosolids are beneficially used in crop production on land located as 

near to the treatment facility as feasible to minimize transportation costs. Transportation 

distances range from essentially zero to several hundred kilometers depending on plant 

location, size and the amount of biosolids or biosolid nutrient content. Phosphorus 

removal will increase both the mass of biosolids and the P content of the biosolids. 

Biosolids application to agricultural land is regulated through the NPDES permit 

of the treatment facility. In many places in the MARB, land application of biosolids is 

based on the N needs of the crop. As with animal manures, biosolids application to meet 

crop N needs results in over application of P and build-up of bio-available P in the soil 

surface. Research during the last two decades has indicated that soil P levels substantially 

in excess of crop needs can cause elevated P concentrations in runoff; particularly from 

critical source areas within fields. As a result, recommendations for application of 

organic nutrient sources, such as manure or biosolids, suggest that applications be limited 

based on P where the risk of loss is moderate to high. This will minimize the opportunity 

for P removed from discharged effluent to be lost in runoff when biosolids are land 

applied. All states now have a tool to estimate the potential for P loss from application of 

manure or biosolids. Nearly all states use a locally adapted version of the Phosphorus Site 

Index (PSI) to estimate P loss risk. Since biosolids currently contain more P relative to N 

than crops require, land application of biosolids should routinely involve an evaluation of 

the risk of P loss using the PSI or another risk assessment tool. 
202

Key Findings and Recommendations

 Sewage treatment plants and industrial dischargers represent a more significant 

source of N and P in the MARB than was originally identified in the Integrated 

Assessment. Tightening effluent limits on large sewage treatment plants together with 

establishing more stringent pretreatment nutrient standards on non-domestic sewer users 

may offer some of the most certain short-term and cost-effective opportunities for 

substantial nutrient reductions, particularly for P, but a full analysis of costs needs to be 

conducted in the context of non-point source reduction costs. Based on these findings, 

the SAB Panel offers the following recommendations. 

• Tighter limits on N and P effluent discharge concentrations for major sewage 

treatment plants, together with concomitant reductions in nutrient discharges from 

non-domestic sewer users, should be considered, following an analysis of the cost 

and technical feasibility for a particular basin. 

• A review of discharge data, including N and P loads, for industrial dischargers 

could identify possible industrial facilities to target for cost-effective reductions. 

• Regulatory authorities should encourage or require sewage treatment plants to 

utilize phosphorus–based biosolids land application rates rather than the nitrogenbased rates in beneficial-use programs. 

 

4.5.9. Ethanol and Water Quality in the MARB 

The production of renewable fuels has been of interest since the 1973 oil price 

shocks, and technologies for the conversion of crops into ethanol and bio-diesel have 

existed since the 1940s. Currently about 99% of renewable transportation fuel produced 

domestically is ethanol from grains and oil crops, primarily corn (Institute for 

Agricultural and Trade Policy [IATP] 2006). This section focuses on the potential water 

quality implications of both ethanol production from corn and its potential production 

from lignocellulosic feedstocks. 

The rapid growth in corn prices is a result of increased ethanol production, 

projected to rise from less to 2 billion gallons in 2001 to more than 19 billions gallons in 

2009, a 950% increase (IATP, 2006). Current estimates are that about 75% of that 

production will be in the nine Upper Mississippi River Corn Belt states (IATP 2006). 

The Food and Agricultural Policy Institute (FAPRI) projects that ethanol production from 

corn will increase from about 6.8 billion gallons in 2007 to over 14 billion gallons by 

2012. Associated with this increase in ethanol production, FAPRI projects an increase in 

corn acreage from about 80 million acres to about 94 million acres in the same time 

period (www.fapri.missouri.edu ). This growth of grain-based ethanol production may 

have major water quality implications for the MARB and the country. 
203

Cellulosic ethanol is an alternative fuel made from a variety of non-food 

feedstocks (such as agricultural residuals like corn stover and cereal straws, industrial 

plant byproducts like saw dust and paper pulp, and crops grown specifically for fuel 

production like switchgrass, Panicum virgatum). By using a variety of regional 

feedstocks for refining cellulosic ethanol, the fuel can be produced in nearly every region 

of the country. Though it requires a more complex refining process, cellulosic ethanol 

produces less impacts on water quality, contains more net energy, and results in lower 

greenhouse emissions than traditional corn-based ethanol (Mclaughlin and Walsh, 1998). 

One of the challenges for wider use of cellulosic ethanol is that the cost of production is 

higher than current prices for corn ethanol and gasoline. Another challenge is that 

technology has not yet developed the fermentation efficiency for conversion of cellulosic 

feedstocks to the level at which it is commercially viable. Contributing to the high cost is 

the need to consolidate enough feedstock close to the plant to produce an adequate supply 

as well as the cost of transporting the heavy and bulky feedstock (Perlack and Turhollow, 

2003). 

Many hope that the heightened interest in biofuels will lead to a more sustainable 

mode of energy production by reducing impacts on water quality, recycling biomass 

residuals and emitting little, if any, greenhouse gases. The vision is that future 

biorefineries will use tailored perennial plants in increasing amounts (Perlack et al., 

2005). Integration of agroenergy plant resources and biorefinery technologies can lead to 

a new manufacturing paradigm (Ragauskas et al., 2006). While these possibilities exist, 

much is unknown concerning how this future might develop and whether it is 

economically and technically viable. 

Water Quality Implications of Projected Grain-based Ethanol Production Levels 

The SAB Panel could find no published estimates of the likely impact of the 

consequences of expanded corn based ethanol production on nutrient flows from the 

MARB. To characterize the short-term potential impact, a set of simple calculations is 

reported in Table 16 that combine acreage projections from the FAPRI baseline for CRP 

and three major field crops in the U.S. with estimates of the per acre nutrient losses from 

these crops (CEAP 2007). The second and third columns in the table report the projected 

nationwide acreage for the years 2007 and 2013 for corn, soybeans, wheat, and CRP and 

the fourth column reports the projected change in acreage for each. As can be seen, the 

FAPRI baseline projects a sizable increase in corn acreage, with that increase coming 

largely from soybeans and the CRP (totals do not add up since other cropland is omitted). 

The fifth column estimates per acre N loss for corn, soybeans, and winter wheat 

based on the sum of waterborne losses reported in the CEAP assessment 

(http://www.nrcs.usda.gov/technical/nri/ceap/croplandreport/ table 36, page 117) for the 

Upper Midwest region. The CEAP report did not estimate N loss from CRP, but for the 

current analysis, losses from CRP are assumed to be 10 % of the average loss from 

cropland. The sixth column reports the estimated change in total N losses due to the 

change in acreage of CRP and each respective crop, with the sum in the bottom row 
204

representing the total projected increase in N loss. By this calculation, N losses 

nationwide could increase by 297 million pounds N /year between 2007 and 2013. 

Implications for nutrient loads to the Gulf of course depend on how much of the 

predicted acreage change will occur in the MARB. Assuming the MARB accounts for 

80% of the change in cropping systems, additional losses of 238 million pounds N / year 

could be expected for the MARB. 

While these estimates are rough and omit numerous factors that could affect the 

nutrient loss from these lands (policy changes, e.g. higher mandates for the ethanol 

content of gasoline, farming practices, energy prices, and climate change) they provide an 

idea of the magnitude of the possible short-term nutrient consequences from increased 

corn-based ethanol production. 
205

Table 16: Estimated changes in N losses from cropping changes predicted by FAPRI from 2006-2013. 

 2006 - 2007 

FAPRI 

Baseline 

(million 

acres) 

2012- 2013 

Acreage 

Projections, 

FAPRI1 

(million 

acres) 

Projected 

Change in 

Acreage2

(million 

acres) 

N Loss 

Estimate per 

acre3

 (lbs./acre) 

Diff. in Total 

N Losses - 

million lbs.4

Corn 78.3 93.7 15.4 28.1 431.6

Soybeans 75.5 67.9 -7.6 17.7 -134.2

Wheat 57.3 58.3 0.9 12.9 11.7

CRP 36.0 30.0 -6.0 2 -12

Total 247.2 249.9 297

1. These projections for U.S. agricultural markets are from the August, 2007 baseline 

http://www.fapri.missouri.edu/outreach/publications/2007/FAPRI_MU_Report_28_07.pdf 

2. This column is the difference between columns 1 and 2. 

3. Per acre estimates of N loss for corn, soybeans, and winter wheat are the sum of waterborne losses 

reported in the CEAP assessment (http://www.nrcs.usda.gov/technical/nri/ceap/croplandreport/ table 36, 

page 117) for the Upper Midwest region. The CEAP report did not estimate N loss from CRP, but for the 

current analysis, losses from CRP are assumed to be 10 % of the average loss from cropland. The CEAP N 

loss rates are based on simulations using the Erosion Productivity Impact Calculator (EPIC) model. The 

CEAP estimates tend to overestimate surface losses and underestimate subsurface losses because EPIC 

does not estimate tile drainage losses that increase the dissolved subsurface loss of nitrate. 

4. The difference in total N losses is computed by multiplying the projected changes in acreage (column 3) 

by the N loss estimate per acre (column 4). 

Impacts on Nutrient Application to Corn 

 In the simple calculations made in Table 16, it was implicitly assumed that N 

application rates will remain unchanged. However, reductions in N application rates have 

been identified as one tool to reduce N loss from corn (CERN, 2000). The level of 

nitrogen application that maximizes farm profits for a given soil and climate is a function 

of price and input costs. Corn price has increased, but fertilizer N costs have also 

skyrocketed in recent years so it is not possible, without further analysis, to determine the 

net effects of these two price trajectories on fertilizer application rates. Further, as 

Sawyer and Randall (2006) point out, simply applying N at economically optimal rates 

will not resolve the issue of nitrate movement from fields in subsurface drainage, for 

nitrate losses occur in corn production systems even when no N is applied. 

 High corn prices associated with market impacts of increased ethanol production 

will make it less profitable for farmers to manage N conservatively. Higher corn prices 

are likely to reinforce the perception that assurance of adequate N is worth the cost, since 

farmers are more likely to be adverse to risks of yield loss when corn prices are high. 

Based on economic optimum yield and historic response to high corn prices by farmers, 
206

$4/bushel corn may tend to increase N application rates to levels where N use efficiency 

is lower. High corn prices also provide a disincentive for cropland retirement or 

conversion to perennials. 

Finally, it is worth noting that a large literature exists on the likely magnitude of 

yield drag associated with continuous corn and other crop rotations. These effects may 

also mean higher fertilization over the levels assumed in the CEAP study used in Table 

16. See Katsvairo and Cox. (2000a and 2000b) and Pikul, Hammack, and Riedell (2005). 

Grain versus Cellulosic Ethanol and Water Quality 

 Cellulosic ethanol produced from perennial grasses, fast-growing woody species, 

manures and other biomass residuals such as corn stover could allow the US to meet 

renewable transportation fuel goals while improving water quality (Mann and Tolbert, 

2000; Perlack et al., 2005). Yet the rapid expansion of grain-based ethanol products may 

be a disincentive to development of perennial crops or crop residual-based ethanol. The 

technology to produce ethanol from cellulosic materials is rapidly improving but is not 

yet operational. The production, storage, and handling infrastructure are in place for 

grain but not for perennial crops or residuals. Cellulosic material is harder to handle and 

only biomass sources such as forestry residuals and corn stover are in sufficient 

abundance to provide reliable supplies. 

Grain-based ethanol producers are interested in the development of technology 

using corn stover and other crop residue as feedstock. Crop residues represent the largest 

potential source of feedstock, projected to be 354 million metric tonne/yr (390 million 

ton/yr). Graham et al. (2007) estimated about 58 million dry metric tonne/yr (64 million 

dry ton/yr) could be removed with soil loss at “tolerable levels” (T) levels, but at 1⁄2 T soil 

loss removals could only be about 18 million metric tonne/yr (19.8 million ton/yr) (at 

1995-2000 corn production levels). However, soil losses could increase 2 to 20 fold and 

still be below T. Therefore harvesting corn stover to keep soil losses just below T would 

result in substantial increases in erosion and associated N and P losses compared to 

current conservation or no-till production. 

 English et al. (2006) proposed that corn stover may be the largest potential source 

of cellulosic materials for ethanol production once cellulosic technologies are cost 

competitive. However, the contribution of returning stover to soil quality and quantity 

has long been recognized. Wilhelm et al. (2004) conclude that corn stover can be 

harvested for ethanol production, but recommendations for removal vary depending on 

regional yield, climatic conditions, and cultural practices. 

 Perennial grasses, including switchgrass and high biomass-producing trees, are 

currently considered the most promising energy crops (Tolbert, 1998; Kurt et al., 1998; 

McLaughlin and Kszos, 2005). Miscanthus and sweet sorghum have also been suggested 

as possible perennial feedstocks. This discussion focuses on switchgrass, which is a 

warm season perennial native prairie grass that produces high biomass in its above 

ground growth and in deep roots. Switchgrass requires some N and P for optimal 
207

production, but less than corn. Switchgrass normally requires two growing seasons to 

become fully productive, but then it can grow for 20 years or more without replanting. 

Thus, either expected profitability from switchgrass production must be large enough to 

overcome early lower yields or an incentive program will be needed to compensate the 

farmer during the two-year transition. As mentioned previously, the transport and storage 

infrastructure needed to handle the large quantities of materials for an ethanol facility will 

need to be developed. 

 The evidence thus far suggests that switchgrass is a more favorable energy crop 

for reducing impacts on the land and climate, however the technology for converting 

switchgrass to ethanol is not yet commercially viable. The fermentation co-product is a 

lignocellulosic material that can be dried and burned to provide part of the energy for the 

facility with net positive energy returns (Farrell et al., 2006). It is very low in nutrients, 

is not suited as a feed amendment, and poses little threat to water quality. If it is grown 

instead of corn on productive soils, N and P losses are expected to be reduced by over 

50% (Chesapeake Bay Program, 2003). Switchgrass will also sequester carbon, increase 

soil organic matter, and improve soil quality through its extensive, deep root system. 

These positive environmental attributes have substantial potential to provide multiple 

revenue streams. Lower production cost, greater net energy production, multiple revenue 

streams and environmental benefits of switchgrass all favor its long-term use as a 

dedicated energy crop. However, the lag in development of fermentation technology and 

the lack of existing infrastructure prevent it from replacing corn as the major ethanol 

feedstock for the near future. 

 Increasing grain prices have increased the relative economic advantage that row 

crops, particularly corn, have over switchgrass. Substantial incentives will be needed 

before farmers would convert row crop land to switchgrass or other perennials at current 

market conditions. Babcock et al. (2007) estimated that the magnitude of subsidies 

would be significant and that conversion of all cropland to switchgrass in a watershed in 

northeastern Iowa would result in an 84%, 83%, 44% and 53% reduction respectively in 

sediment, total phosphorus (TP), nitrate (NO3) and total nitrogen (TN) at the watershed 

outlet compared to existing conditions. Model results also indicated that conversion of 

all cropland in the watershed to continuous corn would increase sediment, TP, NO3, and 

TN from current levels by 23%, 128%, 147% and 150% respectively. They also 

evaluated the impact of growing switchgrass on all Highly Erodible Land (HEL) and 

continuous corn on other cropland. Careful placement of the switchgrass on other 

sensitive landscapes and as a buffer on non-HEL land could provide additional water 

quality benefits.
208

Key Findings and Recommendations 

 Expansion and intensification of corn production to support grain-based ethanol 

production and impacts of ethanol co-products from the animal production sector are 

likely to cause major increases in N and P losses in the MARB. The opportunity still 

exists to make choices that result in a renewable energy strategy that achieves energy 

goals with a reduced impact on the environment. Grain-based ethanol production is 

rapidly expanding, and the SAB Panel’s preliminary calculations demonstrate a 

significant short-run increase in N and P losses to water resulting from current market 

incentives favoring corn. 

 Cellulosic ethanol production can be less environmentally detrimental, but 

current technology and infrastructures do not make it competitive with grain-based 

ethanol. Harvesting corn stover as a feedstock for cellulosic ethanol has water and soil 

quality implications. Switchgrass or other perennial grasses or woody biomass provide 

greater net energy and lower production costs and potentially higher total revenue with 

substantial environmental benefits when compared to corn and could become the 

dominant feedstock if investment, policy, and market conditions do not keep renewable 

energy policy focused on grain feedstocks. Based on these findings, the SAB Panel 

offers the following recommendations regarding biofuel production. 

• Life cycle analysis, examining all impacts to air, water and climate, is needed to 

compare the various feedstocks for ethanol production. 

• Research and development should focus on biofuel production systems that are 

both economically viable and ecologically desirable. 

• If research continues to support the potential of cellulosic materials to meet 

energy and environmental goals, incentives (or the removal of disincentives) 

should be provided to promote ethanol production with more environmentally 

benign feedstocks. 

4.5.10. Integrating Conservation Options 

 The previous sections have described land management and conservation 

practices that can enhance nutrient loss reduction and water quality locally and in the 

Gulf. As discussed, these practices vary, sometimes substantially, in their effectiveness 

among watersheds and subbasins in the MARB. Furthermore, there can be synergistic 

effects on nutrient loss reductions, where combinations of these practices can produce 

more (or less) than the sum of their individual reductions. In evaluating suites of 

management options, it is crucial to determine whether the nutrients that are not released 

to waters are being lost instead to other systems, so that reactive N and P are not actually 
209

removed from the environment, but just redistributed. These facts are an important part of 

the basis for our recommendation that watershed based modeling approaches continue to 

be developed and that they be explicitly used to design optimal land management systems 

within an adaptive management context. As noted in Sections 2.1.9 and 3.4, watershedbased models can be a key source of information for considering alternative sets of 

conservation practices and implementation approaches. Ideally, integrated modeling 

systems would be used to evaluate whether it is more cost-effective to reduce nutrient 

loadings with targeted nutrient management practices on the farm, to subsidize edge-offield buffers in targeted watersheds, to change cropping patterns or to focus financing on 

well-placed off-site freshwater wetlands, or to implement some carefully chosen 

combination of these practices. However, while such models exist and are continuously 

being further improved, there remain limitations of these models in their current state (see 

Sections 2.1.9 and 3.4). 

 In Table 17, we provide a summary of the potential total nitrogen (TN) and 

phosphorus (TP) reduction efficiencies (percent, %) in surface runoff, subsurface flow, 

and tile drainage that can be realized where the various conservation practices could be 

implemented within the MARB. The cost-effectiveness of these measures will vary from 

site to site and with current and future land- and water-use designations. To a large 

extent, these estimates are based on relevant sections of this report and on reports by 

Devlin et al. (2003), Dinnes (2004), and Gitau et al. (2005). Where numeric values for 

reduction efficiency were not included in these reports, relative effects of practices were 

estimated based on expert opinion as negative (- , indicating increased export expected), 

positive (+, indicating reduced export expected), or neutral (±, indicating no significant 

effect expected). 

 
210

Table 17: Potential total nitrogen (TN) and phosphorus (TP) reduction efficiencies (percent change) in 

surface runoff, subsurface flow, and tile drainage. Estimates are average values for a multiple year basis, 

and some of the numbers in this table are based on a very small amount of field information. 

Surface runoff Subsurface 

flow Tile drainage 

Conservation Practice 

TN TP TN TP TN TP 

Nutrient-use efficiency 

Nitrification and urease inhibitors + 1 ± + ± 1 to 

21 2 ± 

Nitrogen timing, rate, and method of application 

 Spring versus Fall application + ± 4 0 to 25 3 ± 10 to 

30%2 ±

 Recommended rate versus. Above 

recommended rate 

28 to 44 2 + 4 + ± 27 to 

50 2 ±

 Subsurface versus. Surface broadcast 50 5 20 6 - ± 16 2 ±

Phosphorus timing, rate and method of 

application 

 Avoid runoff producing rainfall ± 28-57 2 ± + ± + 

 Rate balanced to crop use versus Above 

recommended rate 0 to 25 5 15 to 

47 2 ± 36 2

 ± 25 2

 Subsurface versus Surface broadcast ± 8 to 92 12 ± - ± - 

Manure management 

 Bioenergy, treatment, alternative use, transport 

to nutrient-deficit areas 

+ 6 + 6 + 6 + 6 + 6 + 6

Adoption of comprehensive farm nutrient 

management plan 0-65 5 7 0-45 5 7 + 6 + 6 + 6 + 6

In-field management 

Conservation tillage 

 No-Till versus Conventional tillage 0 to 25 2 5

35 to 

70 2 5 - ± - ± 

Cover crops 50 2 7 to 63 2 + 48 2 13 to 

50 2 + 

Diverse cropping systems and rotations within 

row cropping (78)

25 to 70 2 5 725 to 

88 2 ± ± 52 to 93 2 ± 

Contour plowing and terracing 20 to 55 3 6

30 to 

75 5 7 - ± ± ± 

Drainage management 

 Standard tile drainage versus undrained 25 7 70 7 + ± - - 

 Water table management versus uncontrolled 

drainage - - + + 25 to 

54 2 + 

 Shallow and/or wide versus standard tile 

placement - - + + 39 2 25 to 

42 2

Conversion to CRP 40 2 + 40 2 + 40 to 

97 2 + 

Conversion to perennials crops +60 to 

90 9

+75 to 

95 9 +90 10 + + +

Pasture/grassland management 

 Livestock Exclusion from Streams versus 

Constant Intensive Grazing 

10 to 80 2 732 to 

76 7 911 + 75 2

 ± ± 
211

 Managed Grazing versus Constant Intensive 

Grazing 

-100 to 

80 2 7

0 to 78 2 7 + + + ± 

In-field vegetative buffers 12 to 51 2 5 7

4 to 67 2 5 7 ± ± - - 

Off-site measures 

Sedimentation basins 55 7 65 7 ± ± ± ± 

Riparian buffers Total N Total P 50 to 82 7 9

40 to 

93 7 911

 Nitrate-N Dissolved P 41 to 92 911

28 to 

85 911

+ + ± ±

Wetlands Total P 0 to 79 

7 911

 Dissolved P 

61 to 92 2 7 22 to 

86 7 911

9 to 74 2 + 20 to 

90 911 + 

Tile drainage loss of N and P Subsurface loss of N 

and P 

Surface runoff of N and P 

1

 Relative effects of practices estimated based on expert opinion as negative (- , indicating increased export expected), 

positive (+, indicating reduced export expected), or neutral (±, indicating no significant effect expected). 2

 From Dinnes (2004) report or from SAB Panel report. Values from IA, IL, MO, MN, NE, OH, and OK are included. 3

 From Randall and Sawyer (2005), Nitrogen application Timing, Forms and Methods. p. 73-84. Session 6, 

UMRSHNC (2006) report. 

4

 Increased crop yields afforded with N fertilizer, likely to increase P uptake by crop and removal if harvested. 

5

 From Devlin et al., 2003. 

6

 Improved manure management leads to lower land application and thereby less potential for loss in any pathway. 7

 Values based on data included in Gitau et al. (2005). 

8

 Studies with only corn-soybean systems are not included, although they were included in Dinnes (2004). 9

 Values from Smith et al. 1992. 

10 Values from Randall et al., 1997. 

11 Values are modifications of values in Dinnes (2004) based on values in SAB Panel report. 

Other comments on Table 17: 

• Values for percent nutrient loss reductions are basin-scale averages, derived from edge-of-field and small 

watershed studies and not from widespread implementation. It must be emphasized that there is a great deal of 

site-specificity (spatial and temporal), which results in a wide range in observed conservation practice efficiency. 

• Conservation practices shaded red are likely to have the greatest reduction efficiency on N loss from tile drainage. 

• Conservation practices shaded green are likely to have the greatest reduction efficiencies for N and P loss in 

surface runoff. 

• Conservation practices shaded blue are likely to have the greatest reduction efficiencies for N and P loss in 

subsurface flow. 

• While some of the conservation practices detailed have large local water quality benefits, they may not have a 

major impact on nutrient loss to the Gulf. To help facilitate implementation of practices that reduce nutrient loads 

to the Gulf, local water quality benefits are an essential to MARB-wide adoption of these strategies. 

• Estimates of N and P reductions are only appropriate to areas where a specific conservation practice can be 

implemented. For instance, it would not be effective to implement surface runoff control practices such as 

sedimentation basins on flat lands with no concentrated surface flow of water. To a certain extent, N and P risk 

assessment tools that identify and quantify site vulnerability to N or P loss should be used at a local or field level 

to effectively target practices and to maximize reduction. 

• Implementation of any one of the tabulated conservation practices can positively or negatively influence the 

effectiveness of another. 

• Awareness of the weather forecast in planning any nutrient application or tillage operation is important to 

avoiding rainfall-induced runoff of applied nutrients and erosion. 
212

• The conversion of cropped acres to perennial crops is distinguished from conversion to CRP lands, in that 

perennial crops will include grasses harvested for cellulosic biofuel production, which may receive maintenance or 

low fertilizer N and P inputs. 

• The conversion of lands to CRP and from annual cropping to perennials is expected to decrease N and P loss in 

surface runoff (shaded green), subsurface flow (shaded blue), and tile drainage (shaded red) due to reduced 

fertilizer and manure nutrient inputs and to reduced erosion afforded by increased vegetative cover. 

• Improved N-use efficiency via appropriate timing, rate, and method of application is expected to benefit P loss 

reductions by increasing crop P uptake and removal if harvested. 

 The estimated reduction efficiencies in Table 17 are based on edge-of-field losses 

for studies conducted within the MARB and do not represent expected whole basin 

reductions. These values represent potential reductions only for those areas where the 

particular practices could be implemented and do not address how broadly a practice 

could be applied. The shaded areas indicate those practices expected to have the greatest 

impact on reducing nutrient export from the MARB as a whole: red shading indicates 

conservation practices that translate into N loss reduction in tile drainage, green shading 

is for surface runoff of N and P, and blue shading for nutrient loss in subsurface flow. It 

is clear that where edge-of-field loss estimates are available, there is a large variability in 

reduction efficiencies, which is both temporally and spatially dependent. This inherent 

variability must be recognized when developing conservation or remedial strategies for 

the MARB, in the context of probability of expected outcomes. It is also a key 

component of the conservation premise that there is no “one size fits all” rationale for 

adaptive management. 

 As a complement to the information summarized in Table 17, a second summary 

of the likely environmental benefits is provided in association with the conservation and 

land management. In Table 18 and Table 19, the focus is on the broader contribution 

these practices can have with respect to a wide variety of environmental services 

including local water quality, carbon sequestration in agricultural soils, wildlife habitat, 

biodiversity, general recreational activities, and air pollution. These effects are based on 

the scientific literature and professional judgment, and potential repercussions are 

indicated only as being positive (+) or negative (-) or having no effect (0). 
213

Table 18: Anticipated benefits associated with different agricultural management options. 

Local 

surface 

WQ 

Agricultural 

management 

option 

Reduce 

N load 

to Gulf 

Reduce 

P load 

to Gulf N P & 

seds 

GW 

quality 

Carbon 

sequestration 

Local 

wildlife 

habitat1 

Biodiversity1 

Recreational 

activities 

Air 

pollution 

reduction 

Soil 

quality 

Decrease 

drainage 

intensity 

+ - + - 0 0 + + + 0 0 

Increase 

freshwater 

wetlands 

+ +/? + +/? 0 + + + + - 0 

Forested 

riparian 

buffers 

+ + + + + + + + + + + 

Herbaceous 

riparian 

buffers 

+ + + + + + + + + + + 

Improve 

manure 

mgmt. 

+ + + + + 0 0 0 0 + + 

Increase 

acreage of 

perennials 

+ + + + + + + + + + + 

Increase 

acres of 

farmland 

retired 

+ + + + + + + + + + + 

Reduce 

fertilizer N 

and/ or P 

application 

+ + + + + 0 + + + + 0 

Spring 

fertilizer N 

and/or P 

application 

+ 0 + 0 + 0 0 0 0 0 0 

Expand 

corn-based 

ethanol 

production 

- - - - - - - - - - - 

Expand 

cellulosic 

ethanol 

production 

+ + + + + + + + + + + 

Note: + = will lead to improvements in conditions; - = likely to be further degraded; 0 = will have little 

effect; ? = effect unknown. 
214

Table 19: Anticipated benefits associated with other management options. 

Management 

option 

Reduce 

N load 

to Gulf 

Reduce 

P load 

to Gulf 

Local 

surface 

WQ: 

N 

Local 

surface 

WQ: P 

& seds 

GW 

quality 

Carbon 

sequestration 

Local 

wildlife 

habitat1 

Biodiversity1 

Recreational 

activities 

Air 

pollution 

reduction 

Soil 

quality 

Decrease 

NOx 

emissions + 0 + 0 0 0 0 0 0 + 0 

Reduce 

point source 

loads + + + + 0 0 + + + 0 0 

Reduce 

urban nonpoint source 

loads 

+ + + + + 0 + + + 0 + 

Enhance 

floodplain 

connectivity + + + + 0 + + + + 0 0 

Atchafalaya 

diversion ? ? ? ? 0 0 0 0 ? 0 0 

Increase 

coastal 

wetlands ? + ? ? 0 + + + + 0 + 

Note: + = will lead to improvements in conditions; - = likely to be further degraded; 0 = will have little 

effect; ? = effect unknown. 

 In each of these tables, the effects predicted assume that conservation practices 

are implemented and managed (maintained) as designed to maximize effectiveness and 

life expectancies. Inadequate implementation and maintenance can lead to poor 

performance of such systems. Further, these strategies need to be carefully targeted at an 

appropriate level of intensity and over sufficient time in order to effectively reduce 

nutrient export. 

 Finally, when considering these tables it is important to note there are synergistic 

effects of combinations of conservation practices that result in greater nutrient loss 

reductions than do individual practices (Table 18). For example, N application 

management that minimizes the potential for excess N available to be leached (nutrient 

management, Table 17) should be combined with efforts to reduce the potential off-site 

movement of water (in-field management, Table 17). Conversely, there are potential 

tradeoffs. For example, reduced-till, no-till, and tile drainage can decrease runoff, 

erosion, and P loss but can enhance NO3 nitrate leaching potential. As another example, 

while N-based manure application can be a cost-effective N source to meet crop N needs, 

P may be over applied, increasing the potential for increased runoff and loss of P. 
215

Key Findings and Recommendations 

 A number of conclusions concerning the appropriate use of conservation practices 

can be drawn from these tables. First, there is no “one size fits all” land use or 

conservation practice strategy that will be cost-effective in all locations. Rather, site 

specific and regional optimization of conservation practices and appropriate targeting of 

conservation practices and measures will be needed and will include a broad range of 

alternative practices and land uses such as crop, animal, fertilizer, and drainage 

management measures targeted to appropriate areas. The reduction efficiencies of these 

practices are spatially and temporally variable, making it impossible to assign a specific 

reduction efficiency for any given conservation practice. As information from ongoing 

monitoring of nutrient loss reduction efficiencies becomes available, we will be better be 

able to determine what major factors influence reduction efficiencies. This learning and 

integration of new knowledge is important and will enhance the process of adaptive 

management. 

 Second, practices that are likely to address NGOM hypoxia effectively in tiledrained landscapes can differ markedly from those appropriate in non-tiled lands. 

Further, while there are no-one-size-fits-all strategies, there are some approaches that 

appear particularly promising. For example, inter-seeding of leguminous cover or relay 

crops within corn and other grain rotations can decrease fertilizer N requirements, reduce 

soil profile N at critical loss times of the year, and mine excess soil P. Reconnecting the 

floodplain with managed agricultural lands, by managing hydrology to increase the 

amount of time water is retained on the land (wetland) prior to entering the major fluvial 

systems, should be considered an important part of an adaptive management plan to 

reduce NGOM hypoxia. 

 Third, practices that are likely to be cost-effective in addressing NGOM hypoxia 

may not be the same that yield the highest benefits in other environmental dimensions. 

This has important planning and implementation implications, for it suggests that, when 

considering implementation strategies, the optimal set of conservation practices and sinks 

needs to be considered with respect both to NGOM hypoxia and to the suite of other 

environmental concerns that are likely to vary regionally. 

 Finally, in considering information from the tables and “optimal” sets of 

practices, the principles of adaptive management imply that approaches need to be 

changed and updated with time to maximize overall efficiency. In the process, more 

information can and will be learned about the effectiveness of these practices. This 

information can be used both to improve the performance of water quality models to aid 

in better implementation strategies and directly to improve targeting of conservation 

practices and actions. Based on these findings, the SAB Panel offers the following 

recommendations. 

• There is great temporal and spatial variability in nutrient loss reduction 

efficiencies of the various conservation practices available. Thus, continued, new, 

and enhanced small watershed based studies of suites of conservation practices as 

applied in the real-world are necessary and should be set in a context of research, 
216

monitoring, and demonstration to stakeholders so that progress (or lack thereof) in 

response to management change can be assessed. A variety of response measures 

relevant to different watershed scales and environmental concerns should be 

monitored. These measures should include both performance measures (e.g., 

nutrient loading at sub watershed levels, estimates of carbon sequestered on the 

landscape) and practice-based measures (e.g., number of acres of wetlands 

installed, miles of conservation buffers installed, etc.). 

• To reduce spring nitrate loss from tile drained regions, alternative and more 

complex cropping systems (including perennials) are thought to be the most 

effective method of reducing losses. However, given current constraints in 

cropping systems, the SAB Panel recommends reducing or discontinuing fall N 

application for corn, improved N fertilizer management techniques, use of cover 

crops, wetland establishment, and drainage management where appropriate. 

• For P loss reduction, the Panel again finds that alternative and complex cropping 

systems are most effective. For current cropping systems, the Panel recommends 

that riparian buffer strips, improved P fertilizer and manure management, and 

where appropriate, cover crops be implemented. 

• Where appreciable drainage occurs in the fall and winter, controlled drainage 

could significantly reduce nitrate losses but can be expected to increase surface 

runoff and losses of particulate contaminants. 

• If precision agriculture and controlled release fertilizer technologies are proven to 

provide reductions in losses of N and P to water resources, then incentives should 

be considered to stimulate their adoption. 

• Incentives for conversion to perennials, which have potential future use as 

cellulosic biofuels production, should be established to promote the co-benefit of 

greatly reduced nitrate and P loss from agricultural systems. 

• There should be a focus on conservation practices and implementation strategies 

that appropriately match the nutrient reduction strategies with the goals of 

reducing NGOM hypoxia as well as local/regional environmental goals (carbon 

sequestration, wildlife, air quality, local water quality, etc.). Given the breadth 

and magnitude of these additional environmental goals, these “co-benefits” should 

be incorporated in the planning process. 

• Information on effectiveness and geographic appropriateness of various 

conservation practices and nutrient reduction strategies should be used in 

conjunction with formal models to plan implementation strategies for 

conservation measures that effect a reduction in nutrient loading to the NGOM. 
217

5. Summary of Findings and Recommendations 

 This SAB report provides responses to charge questions in three general areas: 

characterization of hypoxia; characterization of nutrient fate, transport and sources; and 

the scientific basis for goals and management options. In the sections below, charge 

questions are addressed very briefly with references to those sections of this report where 

more detailed science on that particular charge question may be found. 

5.1. Charge Questions on Characterization of Hypoxia 

I. Characterization of Hypoxia – The development, persistence and areal extent of 

hypoxia is thought to result from interactions in physical, chemical and biological 

oceanographic processes along the northern Gulf continental shelf; and changes 

in the Mississippi River basin that affect nutrient loads and fresh water flow. 

A. Address the state-of-the-science and the importance of various processes in the 

formation of hypoxia in the Gulf of Mexico. These issues include: 

i. increased volume and/ or funneling of fresh water discharges from the 

Mississippi River; 

ii. changes in hydrologic or geomorphic processes in the Gulf of Mexico and the 

Mississippi River basin; 

 As discussed in Section 2.1, the hydrologic regime of the Mississippi River and 

spatial distribution and timing of freshwater inputs to the Gulf of Mexico relative to the 

occurrence of energetic currents and waves are critical to vertical mixing intensity, 

stratification, and hypoxia in the Gulf. Alteration of the hydrologic regime of the 

Mississippi and Atchafalaya Rivers from the 1920’s to 1960’s has likely increased the 

residence time of freshwater on the Louisiana-Texas shelf as well as the area of the 

NGOM shelf that is conducive to hypoxia. 

iii. increased nutrient loads due to coastal wetlands losses, upwelling or 

increased loadings from the Mississippi River basin; 

 As discussed in Section 2.1, increased nutrient loadings from the Mississippi 

River basin have triggered hypoxia by stimulating in-situ phytoplankton production of 

labile organic matter in shallow near-shore receiving waters of the Gulf. Nutrients also 

enter this region of the Gulf by advective transport from deeper offshore sources and 

from atmospheric deposition. However, advective imports and atmospheric deposition 

are relatively minor sources of nutrients in comparison with those from the Mississippi 

River basin. The extent to which coastal wetland losses have changed nutrient processing 

and loading to the Gulf of Mexico is a subject of continued study but is largely believed 

to be of secondary importance. 

iv. increased stratification, and seasonal changes in magnitude and spatial 

distribution of stratification and nutrient concentrations in the Gulf; 
218

 As discussed in Section 2.1, increased phytoplankton production, coupled with 

stratification and suppressed vertical mixing associated with fresh water discharge has 

caused hypoxia in bottom waters of the northern Gulf of Mexico. However, historic 

analyses indicate a great deal of variability in seasonal, inter-annual, and decadal scale 

patterns of primary productivity, phytoplankton biomass, and the amounts of freshwater 

and nutrients discharged to the Gulf. Therefore, trends for nutrient-driven eutrophication 

and hypoxia on these time scales have been difficult to interpret. 

v. temporal and spatial changes in nutrient limitation or co-limitation, for 

nitrogen or phosphorus, as significant factors in the development of the 

hypoxic zone; 

 As discussed in Section 2.1.3, studies of waters overlying the hypoxic region of 

the northern Gulf of Mexico indicate that N limitation characterizes offshore waters, but 

inshore productivity appears to be P limited and P and N co-limited. This is particularly 

true from February to May when peak phytoplankton productivity and biomass formation 

coincide with peak freshwater discharge and nutrient loading. Inshore primary 

productivity shifts to an N limited mode during the drier (lower freshwater discharge) 

summer and fall seasons, and there are likely to be periods when both N and P are 

supplied at low levels and co-limit phytoplankton production during the spring to summer 

transition. 

vi. the implications of reduction of phosphorus or nitrogen without concomitant 

reduction of the other. 

 As discussed in Section 2.1, the Panel finds ample evidence to conclude that N 

loading from the Mississippi Atchafalaya River basin is the significant factor driving the 

timing and extent of hypoxia in the northern Gulf of Mexico. However, P supplies also 

play a significant role in controlling primary production. Therefore, as discussed in 

Section 2.1.8, reducing the size of the hypoxic zone requires both N and P discharge 

reductions. 

B. Comment on the state of the science for characterizing the onset, volume, extent and 

duration of the hypoxic zone. 

Section 2.1.9 describes modeling approaches that have been used to characterize 

the onset, volume, extend, and duration of the hypoxic zone. Simple linear and multiple 

regression models that use nutrient loadings to predict hypoxic zone area have been 

constructed. Other models have included some consideration of processes and 

mechanisms. 
219

5.2. Charge Questions on Nutrient Fate, Transport and Sources 

II. Characterization of Nutrient Fate, Transport and Sources: Nutrient loads, 

concentrations, speciation, seasonality and biogeochemical recycling processes have 

been suggested as important causal factors in the development and persistence of 

hypoxia in the Gulf. The Integrated Assessment (CENR 2000) presented information on 

the geographic locations of nutrient loads to the Gulf and the human and natural 

activities that contribute nutrient loadings. 

A. Given the available literature and information (especially since 2000), data and 

models on the loads, fate and transport and effects of nutrients, evaluate the importance 

of various processes in nutrient delivery and effects. These may include: 

i. The pertinent temporal (annual and seasonal) characteristics of nutrient 

loads/fluxes throughout the Mississippi River basin and, ultimately, to the Gulf of 

Mexico. 

 Total annual N flux discharged to the Gulf of Mexico, primarily nitrate-N and 

particulate/organic N, has decreased during the past 25 years, as has the spring (AprilJune) flux. Neither total P nor SRP fluxes show major annual or seasonal trends during 

the same period. 

 As discussed in Section 3.1, the upper Mississippi and Ohio-Tennessee River 

subbasins contribute about 82% of the annual nitrate-N flux, 69% of the TKN flux, and 

58% of the total P flux to the Gulf of Mexico while representing only 31% of the 

drainage area of the MARB. When the upper Mississippi River basin is further divided, 

the subbasin contributing to the upper Mississippi River between Clinton, IA and 

Grafton, IL (only 7% of the drainage area) contributes about 29% of the total annual 

nitrate-N flux to the Gulf. Perhaps more importantly, the upper Mississippi and OhioTennessee River subbasins currently contribute nearly all the spring N flux to the Gulf. 

These subbasins represent the tile-drained, corn-soybean landscape of Iowa, Illinois, 

Indiana, and Ohio and illustrate that corn-soybean agriculture with tile drainage leaks 

considerable N under the current management system. The source of riverine P is more 

diffuse, although these subbasins are also the largest sources of P. 

ii. The ability to determine an accurate mass balance of the nutrient loads 

throughout the basin. 

Estimates of mass balances for nutrient inputs during the period since the 

Integrated Assessment have been recalculated and are discussed in Section 3.2, but the 

research needs described in the Integrated Assessment remain unresolved. Therefore, the 

Panel’s ability to determine an accurate mass balance of nutrient inputs to the MARB is 

limited by the available information and understanding. For example, some components 

of the N mass balance (e.g., denitrification, N2 fixation, manure N, soil N pool processes 

such as mineralization and immobilization) are not measured each year. N2 fixation and 

manure N are the only two of these components that can be estimated. There are too few 

data available for the remaining processes to allow calculations. There also is still a 

disconnect between estimates of inputs to the land (i.e. fertilizer and manure use) and 
220

estimates of the proportion of N and P from those inputs that reach the riverine system 

and contribute to the nutrient flux. Point sources discharge N and P directly to rivers, and 

are estimated by this Panel to contribute about 22% and 34% of the annual riverine N and 

P flux respectively, yet their contributions continue to be estimated from permit limits 

and are not actually measured. Better point-source data are needed to improve mass 

balance estimates of nutrient loads. 

iii. Nutrient transport processes (fate/transport, sources/sinks, transformations, 

etc.) through the basin, the deltaic zone, and into the Gulf. 

 As discussed in Section 3.3, the percentage of annual N and P inputs removed by 

in-stream processes varies by MARB subbasin and ranges from 20 to 55% for N and 20 

to 75% for P based on model estimates. Denitrification can be a significant pathway for 

N removal in small streams during low flow, warm periods, thereby enhancing local 

water quality. However, most nitrate-N is exported to the Gulf during high flows in the 

period from January to June, when denitrification is not an effective removal process. 

Although current estimates of denitrification rates in coastal wetlands are higher than the 

estimates used in the Integrated Assessment, current studies still conclude that river 

diversions to coastal wetlands would remove only small amounts of nutrients relative to 

the total fluxes. However, better estimates of nutrient and organic matter loss rates 

(denitrification; long-term burial of C, N, and P; and plant uptake) are needed to better 

understand observed differences between wetland inputs and outputs in coastal areas. 

B. Given the available literature and information (especially since 2000) on nutrient 

sources and delivery within and from the basin, evaluate capabilities to: 

i. Predict nutrient delivery to the Gulf, using currently available scientific tools 

and models; and 

ii. route nutrients from their various sources and account for the transport 

processes throughout the basin and deltaic zone, using currently available 

scientific tools and models. 

 

 In Section 3.4, the SAB Panel singled out three models for discussion: 

SPARROW, SWAT, and IBIS/THMB. Each is capable of N and P load estimation on 

the scale of the MARB, yet each has strengths and weaknesses requiring further 

development. The uncertainty of results from each model reflects the uncertainty of the 

model structure and algorithms, as well as that propagated by the input data, user 

parameterization, the calibration process, other user-defined conditions, and the skill of 

the model user. Even though the capability to predict and route nutrients throughout the 

MARB has improved since the Integrated Assessment, future adaptive management will 

require a smooth interface between watershed, economic, and Gulf of Mexico hypoxia 

models that will allow resource managers the capability to assess the effects of policy 

decisions and management practices on the sources, fate, and transport of nutrients from 

the MARB to the Gulf of Mexico. 
221

5.3. Charge Questions on Goals and Management Options 

III. Scientific Basis for Goals and Management Options. The Task Force has stated 

goals of reducing the 5-year running average areal extent of the Gulf of Mexico hypoxic 

zone to less than 5,000 square kilometers by the year 2015, improving water quality 

within the basin and protecting the communities and economic conditions within the 

basin. Additionally, nutrient loads from various sources in the Mississippi River basin 

have been suggested as the major driver for the formation, extent and duration of the 

Gulf hypoxic zone. 

A. Are these goals supported by present scientific knowledge and understanding of the 

hypoxic zone, nutrient loads, fate and transport, sources and control options? 

 The SAB Panel affirms the major findings of the Integrated Assessment. 

Although the 5,000 km2

 target remains a reasonable endpoint for continued use in an 

adaptive management context; it may no longer be possible to achieve this goal by 2015. 

Accordingly, it is even more important to proceed in a directionally correct fashion to 

manage factors affecting hypoxia than to wait for greater precision in setting the goal for 

the size of the zone. 

i. Based on the current state-of- the-science, should the reduction goal for the size 

of the hypoxia zone be revised? 

 No. As discussed in the Executive Summary, it is more important to begin to 

move in a directionally correct fashion than to refine the goal for the exact size of the 

hypoxic zone. 

ii. Based on the current state-of-the-science, can the areal extent of Gulf hypoxia 

be reduced while also protecting water quality and social welfare in the basin? 

 Social welfare can be protected by choosing policies that incorporate targeting, 

provide economic incentives and maximize co-benefits. As discussed in Section 4.3, 

improvements in large-scale integrated economic and bio-physical models are needed to 

better capture system-wide response and effects. 

B. Based on the current state-of- the-science, what level of reduction in causal agents 

(nutrients/discharge) will be needed to achieve the current reduction goal for the size of 

the hypoxic zone?

 As discussed in Section 4.2, to reduce the size of the hypoxic zone, the SAB Panel 

recommends an adaptive management approach targeting at least a 45% reduction in 

discharges of total N and total P from the 1980 – 1996 fluxes. 

C. Given the available literature and information (especially since 2000) on 

technologies and practices to reduce nutrient loss from agriculture, runoff from other 
222

non-point sources and point source discharges, discuss options (and combinations of 

options) for reducing nutrient flux in terms of cost, feasibility and any other social 

welfare considerations.

 In general, the social costs of reducing nutrients will vary widely with the policy 

chosen, hence overall cost-effectiveness is largely a function of policy. Policies that 

target and provide economic incentives are essential to minimize costs. A wide range of 

policy options are discussed in Section 4.4, while management options are covered 

extensively in Section 4.5. 

These options may include: 

i. the most effective agricultural practices, considering maintenance of soil 

sustainability and avoiding unintended negative environmental consequences. 

 The cost and reduction efficency rankings of agricultural management practices 

will vary by site and region, historic land use and management, depending on crops 

grown, local soil conditions, distance to waterway, field slopes and configuration, 

presence of buffers, drainage structures and so forth. Table 16 in Section 4.5.10 provides 

the SAB Panel’s summary of the evidence comparing the relative effectiveness of 

nutrient (N and P) reduction options in agriculture. Section 4.5.6 discusses management 

options for in-field nutrients. A targeted and adaptive management framework will 

maximize local and regional water quality benefits in the MARB and Gulf. 

ii. the most effective actions for other non-point sources 

 As discussed in Section 4.5.7, there are significant policy opportunities to reduce 

atmospheric deposition of N, however a detailed examination of air pollution control 

policy options was beyond the SAB Panel’s scope. Nonetheless, the Panel strenuously 

recommends incorporating water quality benefits and effects on hypoxia in air pollution 

control decisions. 

iii. the most effective technologies for industrial and municipal point sources. 

 As discussed in Section 4.5.8, a targeted permit by permit approach to industrial 

point source discharges could yield significant opportunities for nutrient (N and P) 

reduction since frequently a limited number of permitted facilities are responsible for a 

large part of the N and P loads. Municipal point sources are also discussed in Section 

4.5.8 where the SAB Panel recommends an analysis to assess the cost and feasibility of 

tightening limits on N and P concentrations in discharges for large sewage treatment 

plants. 

In all three areas, please address research and information gaps (expanded monitoring, 

documentation of sources and management practices, effects of practices, further model 

development and validation, etc.) that should be addressed prior to the next 5-year 

review. 
223

 Recommendations for monitoring and research are found in nearly every section 

of the report and are included below in the summary of the SAB Panel’s 

recommendations. 

5.4. Conclusion 

 This report constitutes the SAB Panel’s response to charge questions posed by the 

EPA Office of Water. This Advisory reaffirms the major findings of the Integrated 

Assessment, while pointing out the need for economic incentives to encourage 

conservation in the Mississippi Atchafalaya River basin. Although the science has 

grown, actions to control hypoxia have lagged. The SAB urges the EPA and other 

agencies to utilize the recommendations of this Advisory and move ahead with 

implementing programs, strategies and policies to reduce the size of the hypoxic zone and 

improve water quality in the Mississippi Atchafalaya River basin. 

 Most of the research and monitoring needs identified in the Integrated Assessment

have not been met, and fewer rivers and streams are monitored today than in 2000. The 

majority of monitoring recommendations in the Integrated Assessment remain relevant 

and should be heeded, specifically the CENR’s call to improve and expand monitoring of 

the temporal and spatial extent of hypoxia and the processes controlling its formation; the 

flux of nutrients, carbon, and other constituents from non-point sources throughout the 

MARB and to the NGOM; and measured (rather than estimated) nitrogen and phosphorus 

fluxes from municipal and industrial point sources. Echoing the CENR, the SAB Panel 

affirms the need for research on the ecological effects of hypoxia; watershed nutrient 

dynamics; effects of different agricultural practices on nutrient losses from land, 

particularly at the small watershed scale; nutrient cycling and carbon dynamics; longterm changes in hydrology and climate; and economic and social impacts of hypoxia. A 

suite of models is needed to simulate the processes and linkages that regulate the onset, 

duration and extent of hypoxia. Emerging coastal ocean observation and prediction 

systems should be encouraged to monitor dissolved oxygen and other physical and 

biogeochemical parameters needed to continue improving hypoxia models. 

 Although there are over 90 recommendations in this report, the following major 

recommendations reflect the SAB Panel’s consideration of the new science that has 

emerged since the Integrated Assessment. 

 To advance the science characterizing hypoxia and its causes, the SAB Panel 

finds that research is needed to: 

• collect and analyze additional sediment core data needed to develop a better 

understanding of spatial and temporal trends in hypoxia; 

• investigate freshwater plume dispersal, vertical mixing processes and 

stratification over the Louisiana-Texas continental shelf and Mississippi Sound, 
224

and use three-dimensional hydrodynamic models to study the consequences of 

past and future flow diversions to NGOM distributaries; 

• advance the understanding of biogeochemical and transport processes affecting 

the load of biologically available nutrients and organic matter to the Gulf of 

Mexico, and develop a suite of models that integrate physics and 

biogeochemistry; 

• elucidate the role of P relative to N in regulating phytoplankton production in 

various zones and seasons, and investigate the linkages between inshore primary 

production, offshore production, and the fate of carbon produced in each zone; 

• improve models that characterize the onset, volume, extent, and duration of the 

hypoxic zone, and develop modeling capability to capture the importance of P, N, 

and P-N interactions in hypoxia formation. 

 With respect to advancing the science on sources, fate and transport of nutrients, 

the SAB Panel finds that research is needed to: 

• develop models to simulate fluvial processes and estimate N and P transfer to 

stream channels under different management scenarios; 

• improve the understanding of temporal and seasonal nutrient fluxes and develop 

nutrient, sediment, and organic matter budgets within the MARB; 

 To enhance the scientific basis for implementation of management options, the 

SAB Panel finds that research is needed to: 

• examine the efficacy of dual nutrient control practices; 

• determine the extent, pattern, and intensity of agricultural drainage as well as 

opportunities to reduce nutrient discharge by improving drainage management; 

• integrate monitoring, modeling, experimental results, and ongoing management 

into an improved conceptual understanding of how the forces at key management 

scales influence the formation of the hypoxia zone; and 

• develop integrated economic and watershed models to fully assess costs and 

benefits, including co-benefits, of various management options. 

 To reduce the size of the hypoxic zone, the SAB Panel recommends at least a 

45% reduction in N accompanied by a comparable reduction in P. The Panel found five 

areas that offer the most significant opportunities for N and P reductions: 

• promotion of environmentally sustainable approaches to biofuel production and 

associated cropping systems (e.g. perennials). 
225

• improved management of nutrients by emphasizing infield nutrient management 

efficiency and effectiveness to reduce losses; 

• construction and restoration of wetlands, as well as criteria for targeting those 

wetlands that may have a higher priority for reducing nutrient losses; 

• introduction of tighter N and P limits on municipal point sources; and 

• improved targeting of conservation buffers, including riparian buffers, filter strips 

and grassed waterways, to control surface-borne nutrients. 
226

References 

Ackerman, F., Biewald, B., White, D., Woolf, T., and Moomaw, W., 1999, Grandfathering and coal plant emissions—The cost of cleaning up the Clean Air Act: 

Energy Policy, v. 27, p. 929–940. 

Adviento-Borbe, M.A.A., Haddix, M.L., Binder, D.L., Walters, D.T., and Dobermann, 

A., 2007, Soil greenhouse gas fluxes and global warming potential in four highyielding maize systems: Global Change Biology, in press, listed online as Accepted 

Articles at: http://www.blackwell-synergy.com/doi/abs/10.1111/j.1365-

2486.2007.01421.x, last accessed August 15, 2007. 

Aillery, M., Gollehon, N., Johansson, R., Kaplan, J., Key, N., and Ribaudo, M., 2005, 

Managing manure to improve air and water quality: Washington, D.C., U.S. 

Government Printing Office, U.S. Department of Agriculture, Economic Research 

Service, Economic Research Report 9, 65 p., available online at:

http://www.ers.usda.gov/publications/err9/err9.pdf

Alexander, R.B., Boyer, E.W., Smith, R.A., Schwarz, G.E., and Moore, R.B., 2007a, The 

role of headwater streams in downstream water quality: Journal of the American 

Water Resources Association, v. 43, no. 1, p. 41–59. 

Alexander, R.B., Elliott, A.H., Shankar, U., and McBride, G.B., 2002a, Estimating the 

sources and transport of nutrients in the Waikato River basin, New Zealand: Water 

Resources Research, v. 38, p. 1268–1290. 

Alexander, R.B., Johnes, P.J., Boyer, E.W., and Smith, R.A., 2002b, A comparison of 

models for estimating the riverine export of nitrogen from large watersheds: 

Biogeochemistry, v. 57/58, p. 295–339. 

Alexander, R.B., Slack, J.R., Ludtke, A.S., Fitzgerald, K.K., and Schertz, T.L., 1998, 

Data from selected U.S. Geological Survey national stream water-quality monitoring 

networks: Water Resources Research, v. 34, no. 9, p. 2401–2405. 

Alexander, R.B., and Smith, R.A., 2006, Trends in the nutrient enrichment of U.S. rivers 

during the late 20th century and their relation to changes in probable stream trophic 

conditions: Limnology and Oceanography, v. 51, p. 639–654. 

Alexander, R.B., Smith, R.A., and Schwarz, G.E., 2000, Effect of stream channel size on 

the delivery of nitrogen to the Gulf of Mexico: Nature, v. 403, p. 758–761. 

Alexander, R.B., Smith, R.A., and Schwarz, G.E., 2004, Estimates of diffuse phosphorus 

sources in surface waters of the United States using a spatially referenced watershed 

model: Water Science and Technology, v. 49, no. 3, p. 1–10. 

Alexander, R.B., Smith, R.A., Schwarz, G.E., Boyer, E.W., Nolan, J.V., and Brakebill, 

J.W., in press, Differences in phosphorus and nitrogen delivery to the Gulf of 

Mexico from the Mississippi River basin: Environmental Science and Technology, 

(In press), (accepted for publication 10/30/2007). 

Alexander, R.B., Smith, R.A. Schwarz, G.E., Preston, S.D., Brakebill, J.W., Srinivasan, 

R., and Pacheco, P.A., 2001, Atmospheric nitrogen flux from the watersheds of 

major estuaries of the United States: An application of the SPARROW watershed 
227

model, in Valigura, R., Alexander, R., Castro, M., Meyers, T., Paerl, H., Stacey, P., 

and Turner, R.E., eds., Nitrogen loading in coastal water bodies—An atmospheric 

perspective: American Geophysical Union Monograph 57, p. 119–170. 

Aller, R.C., 1998, Mobile deltaic and continental shelf muds as sub-oxic, fluidized bed 

reactors: Marine Chemistry, v. 61, p. 143–155. 

Aller, R.C., Heilbrun, C., Panzeca, C., Zhu, Z.-B., and Baltzer, F., 2004. Coupling 

between sedimentary dynamics, early diagenetic processes, and biogeochemical 

cycling in the Amazon-Guianas mobile mud belt: Coastal French Guiana: Marine 

Geology, v. 208, p. 331-360. 

Ammerman, J.W., and Sylvan, J.B., 2004, Phosphorus limitation of phytoplankton 

growth in the Mississippi River plume—A case for dual nutrient control?: EOS 

Transactions AGU, v. 85, no. 47, Fall Meeting Supplement, Abstract OS11B-07.

Anand, S., Mankin, K.R., McVay, K.A., Janssen, K.A., Barnes, P.L., and Pierzynski, 

G.M., 2007, Calibration and validation of ADAPT and SWAT for field-scale runoff 

prediction: Journal of the American Water Resources Association, v. 43, no. 4, p. 

899-910. 

Anderson, C.J., Nairn, R.W., and Mitsch, W.J., 2005, Temporal and spatial development 

of surface soil conditions at two created riverine marshes: Journal of Environmental 

Quality, v. 34, p. 2072–2081. 

Anderson, D.M., and Garrison, D.J., eds., 1997, The ecology and oceanography of 

harmful algal blooms: American Society of Limnology and Oceanography Special 

Issue, v. 42, no. 5, p. 1009–1305. 

Andraski, T.W., and Bundy, L.G., 2003, Relationships between phosphorus levels in soil 

and in runoff from corn production systems: Journal of Environmental Quality, v. 32, 

p. 310–316. 

Arnold, J.G., Srinivasan, R., Muttiah, R.S., and Allen, P.M., 1999, Continental scale 

simulation of the hydrologic balance: Journal of the American Water Resources 

Association, v. 35, no. 5, p. 1037–1051. 

Arnold, J.G., Srinivasan, R., Muttiah, R.S., and Williams, J.R., 1998, Large area 

hydrologic modeling and assessment—Part I, Model development: Journal of the 

American Water Resources Association, v. 34, no. 1, p. 73–89. 

Atwood, J.D., Benson, V.W., Srinivasan, R., Walker, C., and Schmid, E., 2001, 

Simulated nitrogen loading from corn, sorghum, and soybean production in the 

Upper Mississippi Valley, in Stott, D.E., Mohtar, R.H., and Steinhardt, G.C., eds., 

Sustaining the Global Farm, 10th International Soil Conservation Organization 

Meeting, Purdue University, IN, May 24–29, 1999, p. 344–348. 

Aulenbach, B.T., and Hooper, R.P., 2006, The composite method—An improved method 

for stream-water solute load estimation: Hydrological Processes, v. 20, p. 3029–

3047. 

Aulenbach, B.T., Buxton, H.T., Battaglin, W.A., and Coupe, R.H., 2007, Streamflow and 

nutrient fluxes of the Mississippi-Atchafalaya River basin and subbasins for the 
228

period of record through 2005: U.S. Geological Survey Open-File Report 2007-1080, 

available online at: http://toxics.usgs.gov/pubs/of-2007-1080/index.html. 

Babcock, B.A., Gassman, P.W., Jha, M., and Kling, C.L., 2007, Adoption subsidies and 

environmental impacts of alternative energy crops: Iowa State University, Center for 

Agricultural and Rural Development (CARD) Briefing Paper 07-BP 50, 15 p, 

available on line at: 

http://www.card.iastate.edu/publications/DBS/PDFFiles/07bp50.pdf. 

Baker, J.L., David, M.B., and Lemke, D.W., in press, Understanding nutrient fate and 

transport, including the importance of hydrology in determining losses, and potential 

implications on management systems to reduce those losses, in Proceedings of Gulf 

Hypoxia and Local Water quality Concerns Workshop, Ames, IA, September 26-28, 

2005. (in press) 

Baker, J.L., Melvin, S.W., Lemke, D.W., Lawlor, P.A., Crumpton, W.G., and Helmers, 

M.J., 2004, Subsurface drainage in Iowa and the water quality benefits and problem, 

in Cooke, R., ed., Proceedings of the Eighth International Drainage Symposium, 

Sacramento, CA, March 21, 2004, ASAE Pub #701P0304, p. 39–50. 

Baker, J.L., Mickelson, S.K., and Crumpton, W.G., 1997, Integrated crop management 

and off-site movement of nutrients and pesticides, in Hatfield, J.C., Buhler, D.B., and 

Stewart, B.A., eds., Weed biology, soil management, and weed management—

Advances in soil science: Boca Raton, CA, CRC Press, p. 135–160. 

Baker, D.B., and Richards, P.R., 2002, Phosphorus budgets and riverine phosphorus 

export in northwestern Ohio watersheds: Journal of Environmental Quality, v. 31, 

p. 96–108. 

Baltz, D.M., Hiram, W.L., Rossignol, P.A., Chesney, E.J., and Switzer, T.S., 2006, A 

qualitative assessment of the relative effects of bycatch reduction of fisheries and 

hypoxia on coastal nekton communities in the Gulf of Mexico: Paper presented at 

Hypoxia Effects on Living Resources in the Gulf of Mexico, September 25–26, 

2006: New Orleans, Louisiana, Tulane University, sponsored by National Oceanic 

and Atmospheric Administration Center for Sponsored Coastal Ocean Research. 

Barker, D.W., Sawyer, J.E., and Al-Kaisi, M.M., 2006a, Assessment of the amino sugarnitrogen test on Iowa soils—I. Evaluation of soil sampling and corn management 

practices: Agronomy Journal, v. 98, p. 1345–1351. 

Barker, D.W., Sawyer, J.E., and Al-Kaisi, M.M., 2006b, Assessment of the amino sugarnitrogen test on Iowa soils—II. Field correlation and calibration: Agronomy Journal, 

v. 98, p. 1352–1358. 

Batie, S.S., Gilliam, J.W., Groffman, P.M., Hallberg, G.R., Hamilton, N.D., Larson, 

W.E., Lee, L.K., Nowak, P.J., Renard, K.G., Rominger, R.E., Stewart, B.A., Tanji, 

K.K., Van Schilfgaarde, J., Wagenet, R.J., and Young, D.L., 1993, Soil and water 

quality: An agenda for agriculture: National Academy of Sciences, Board of 

Agriculture, 278 p. 

Battaglin, W., 2006, Streamflow and nitrogen, phosphorus, and silica flux at selected 

sites in the Mississippi River basin, 1980–2005, presented at Science Symposium: 
229

Sources, Transport and Fate of Nutrients in the Mississippi and Atchafalaya River 

basins, November 7–9, 2006, Minneapolis, MN. 

Baumol, W., and Oates, W., 1988, The Theory of Environmental Policy, 2nd edition: 

Cambridge University Press, Cambridge, UK, 299 p. 

Beegle, D.B., 2005, Assessing soil phosphorus for crop production by soil testing, in

Sims, J.T., and Sharpley, A.N., eds., Phosphorus: Agriculture and the Environment, 

Madison, WI, American Society of Agronomy Monograph Series No. 46, p. 123-

144. 

Belabbassi, L., 2006, Examination of the relationship of river water to occurrences of 

bottom water with reduced oxygen concentrations in the northern Gulf of Mexico: 

College Station, Texas, Texas A&M University, Ph.D. thesis, xii + 119 p. 

Benner, R., and Opsahl, S., 2001, Molecular indicators of the sources and transformations 

of dissolved organic matter in the Mississippi River plume: Organic Geochemistry, 

v. 32, p. 597–611. 

Bermudez, M., and Mallarino, A.P., 2007, Impacts of varable-rate phosphorus 

fertilization based on dense grid soil sampling on soil-test phosphorus and grain 

yield of corn and soybean: Agronomy Journal, v. 99, p. 822-832. 

Bernot, M.J., and Dodds, W.K., 2005, Nitrogen retention, removal, and saturation in lotic 

ecosystems: Ecosystems, v. 8, p. 442-453. 

Besiktepe, S.T., Lermusiaux, P.F.J., and Robinson, A.R., 2003, Coupled physical and 

biogeochemical data-driven simulations of Massachusetts Bay in late summer: Realtime and postcruise data assimilation: Journal of Marine Systems, v. 40-41, p. 171-

212.

Beven, K.J., 2001, Rainfall-runoff modeling: The primer: Wiley, Chichester, UK, 360 p. 

Bharati, L., Lee, K.H., Isenhart, T.M., and Schultz, R.C., 2002, Soil-water infiltration 

under crops, pasture and established riparian buffer in Midwestern USA: 

Agroforestry Systems, v. 56, p. 249–257. 

Bianchi, T..S., Allison, M.A., Canuel, E.A., Corbett, D.R., McKee, B.A., Sampere, T.P., 

Wakeham, S.G., Waterson, E., 2006, Rapid export of organic matter to the 

Mississippi canyon, Mississippi: EOS, Transactions of the American Geophysical 

Union, v. 87, no. 50, p. 572–574. 

Bianchi, T.S., Filley, T., Dria, K., and Hatcher, P.G., 2004, Temporal variability in 

sources of dissolved organic carbon in the lower Mississippi River: Geochimica et 

Cosmochimica Acta, v.68, no. 5, p. 959–967. 

Bianchi, T.S., Galler, J.J., and Allison, M.A.,2007, Hydrodynamic sorting and transport 

of terrestrially derived organic carbon in sediments of the Mississippi and 

Atchafalaya Rivers: Estuarine, Coastal and Shelf Science, v. 73, nos. 1–2, p 211–

222. 

Bianchi, T.S., Mitra, S., and McKee, B., 2002, Sources of terrestrially-derived carbon in 

the lower Mississippi River and Louisiana shelf—Implications for differential 
230

sedimentation and transport at the coastal margin: Marine Chemistry, v. 77, p. 211–

223.

Bierman, V.J., Jr., Hinz, S.C., Zhu, D.-W., Wiseman, W.J., Jr., Rabalais, N.N., and 

Turner, R.E., 1994, A preliminary mass balance model of primary productivity and 

dissolved oxygen in the Mississippi River Plume/Inner Gulf Shelf region: Estuaries, 

v. 17, no. 4, p. 886–899. 

Blaylock, A.D., 2006, Review of enhanced-efficiency nitrogen fertilizers: in Proceedings 

of Southern Plant Nutrient Management Conference, Olive Branch, MS, October 3-

4, 2006, p. 4-10. 

Blomqvist, S., Gunnars, A., and Elmgren, R., 2004, Why the limiting nutrient differs 

between temperate coastal seas and freshwater lakes—A matter of salt: Limnology & 

Oceanography, v. 49, p. 2236–2241.

Bode, A., and Dortch, Q., 1996, Uptake and regeneration of inorganic nitrogen in coastal 

waters influenced by the Mississippi River—Spatial and seasonal variations: Journal 

of Plankton Research. v. 18, no. 12, p. 2251–2268. 

Boesch, D.F., 2002, Challenges and opportunities for science in reducing nutrient overenrichment of coastal ecosystems: Estuaries, v. 25, p. 744–758. 

Boesch, D.F., 2003, Continental shelf hypoxia: Some compelling answers: Comments on 

“Continental shelf hypoxia: Some nagging questions”: Gulf of Mexico Science, v. 

21, no. 2, p. 202-205. Available on line at: 

http://goms.disl.org/tocpages/december2003vol21no2.htm

Boicourt, W.C., 1992, Influences of circulation processes on dissolved oxygen in the 

Chesapeake Bay, in Smith, D.E., Leffler, M., and Mackiernan, G., eds., Oxygen 

dynamics in the Chesapeake Bay—A synthesis of recent research: College Park, 

MD, Maryland Sea Grant College, p. 7–79. 

Booth, M.S., and Campbell, C., 2007, Spring nitrate flux in the Mississippi River basin: 

A landscape model with conservation applications: Environmental Science and 

Technology, v. 41, no. 15, p. 5410-5418. 

Boyer, E.W.,Goodale, C.L., Jaworski, N.A., and Howarth, R.W., 2002, Anthropogenic 

nitrogen sources and relationships to riverine nitrogen export in the northeastern 

U.S.A.: Biogeochemistry, v. 57/58, p. 137–169. 

Boyer, E.W., and Howarth, R.W., in press, Nitrogen fluxes from rivers to the coastal 

oceans, in Capone, D., Mulholland, M., and Carpenter, E., eds., Nitrogen in the 

marine environment, 2nd ed.: New York, NY, Academic Press, (in press). 

Boynton, W.R., and Kemp, W.M., 2000, Influence of river flow and nutrient loads on 

selected ecosystem processes: A synthesis of Chesapeake Bay data, in Hobbie, J.E., 

ed., Estuarine science—A synthetic approach to research and practice: Washington, 

D.C., Island Press, p. 269–298.

Bradley, M.J., and Jones, B.M., 2002, Emissions—Developing advanced energy and 

transportation technologies: AMBIO, A Journal of the Human Environment, v. 31, 

no. 2, p. 141–149. 
231

Bratkovich, A., Dinnel, S.P., and Goolsby, D.A., 1994, Variability and prediction of 

freshwater and nitrate fluxes for the Louisiana-Texas shelf—Mississippi and 

Atchafalaya river source functions: Estuaries, v. 17, p. 766–778. 

Breetz, H., Fisher-Vander, K., Garzon, L., Jacobs, H., Droetz, K., and Terry, R., 2004, 

Water quality trading and offset initiatives in the U.S.—A comprehensive survey: 

Dartmouth College, Hanover, NH, 337 p. Available on line at: 

http://www.dartmouth.edu/~kfv/waterqualitytradingdatabase.pdf.

Brezonik, P.L., Bierman, V.J., Jr., Alexander, R., Anderson, J., Barko, J., Dortch, M., 

Hatch, L., Hitchcock, G.L., Keeney, D., Mulla, D., Smith, B., Walker, C., Whitledge, 

T., and Wiseman, W.J., Jr., 1999, Effects of reducing nutrient loads to surface waters 

within the Mississippi River basin and the Gulf of Mexico: Topic 4 report for the 

integrated assessment of hypoxia in the Gulf of Mexico: Silver Spring, MD, National 

Oceanic and Atmospheric Administration Coastal Ocean Program Decision Analysis 

Series No. 18, 158 p. Available on line at: 

http://oceanservice.noaa.gov/products/hypox_t4final.pdf. 

Bridgham, S.D., Johnston, C.A., Schubauer-Berigan, J.P., and Wesihampel, P., 2001, 

Phosphorus sorption dynamics in soils and coupling with surface and pore water in 

riverine wetlands: Soil Sciences Society of America Journal, v. 65, p. 577–588. 

Broshears, R.E., Clark, G.M., and Jobson, H., 2001, Simulation of stream discharge and 

transport of nitrate and selected herbicides in the Mississippi River Basin: 

Hydrological Processes, v. 15, p. 1157–1167. 

Brouwer, M., 2006, Changes in gene and protein expression and reproduction in grass 

shrimp, Palaemonetes pugio, exposed to chronic hypoxia: Paper presented at 

Hypoxia Effects on Living Resources in the Gulf of Mexico, September 25–26, 

2006, Tulane University, New Orleans, Louisiana, sponsored by National Oceanic 

and Atmospheric Administration Center for Sponsored Coastal Ocean Research. 

Bruland, G.L., and Richardson, C.J., 2006, An assessment of the phosphorus retention 

capacity of wetlands in the Painter Creek Watershed, Minnesota, USA: Water, Air, 

and Soil Pollution, v. 171, p. 169–184. 

Bundy, L.G., 1998, Corn fertilization: University of Wisconsin Cooperative Extension 

Service Publication AA3340. 

Cai, W.J., and Lohrenz, S.E., 2005, Carbon, nitrogen, and phosphorus fluxes from the 

Mississippi River and the transformation and fate of biological elements in the river 

plume and the adjacent margin, in Liu, K.K., Atkinson, L., Quinones, R., TalaueMcManus, L., eds., Carbon and nutrient fluxes in continental margins—A global 

synthesis: NY, Springer-Verlag.

Cai, W.J., 2003, Riverine inorganic carbon flux and rate of biological uptake in the 

Mississippi River plume: Geophysical Research Letters, v. 30, no. 2, 

doi:10.1029/2002GL016312.

Caraco, N., Cole, J.J., and Likens, G.E., 1990, A comparison of phosphorus 

immobilization in sediments of freshwater and coastal marine systems: 

Biogeochemistry, v. 9, p. 277–290. 
232

Caraco, N.F., Cole, J.J., and Likens, G.E., 1989, Evidence for sulfate controlled 

phosphorus release from sediments of aquatic systems: Nature, v. 341, p. 316–318. 

Carter, G.S., Gregg, M.C., and Lien, R.C., 2005, Internal waves, solitary like waves, and 

mixing on the Monterey Bay shelf: Continental Shelf Research, v. 25, p. 1499–1520. 

CENR, 2000, Integrated assessment of hypoxia in the northern Gulf of Mexico: National 

Science and Technology Council, Committee on Environmental and Natural 

Resources, May 2000, 66 p. Available on line at: 

http://oceanservice.noaa.gov/products/hypox_final.pdf. 

Cerco, C.F., and Cole, T., 1993, Three-dimensional eutrophication model of Chesapeake 

Bay: Journal of Environmental Engineering, v. 119, no. 6, p. 1006-1025. 

Chen, N., Bianchi, T.S., and McKee, B.A., 2005, Early diagnensis of chloropigment 

biomarkers in the lower Mississippi River and Louisiana shelf—Implications for 

carbon cycling in a river-dominated margin: Marine Chemistry, v. 93, p. 159–177.

Chen, N., Bianchi, T.S., McKee, B.A., and Bland, J.M., 2001, Historical trends of 

hypoxia on the Louisiana shelf—Applications of pigments as biomarkers: Organic 

Geochemistry, v. 32, p. 543–561.

Chen, R.F., and Gardner, G.B., 2004, High resolution measurements of chromophoric 

dissolved organic matter in the Mississippi and Atchafalaya river plume regions: 

Marine Chemistry, v. 89, p. 103–125. 

Chen, X., Lohrenz, S.E., and Wiesenburg, D.A., 2000, Distribution and controlling 

mechanisms of primary production over the Louisiana-Texas continental shelf: 

Journal of Marine Systems, v. 25, p. 179–207. 

Chesapeake Bay Commission, 2004, Cost effective strategies for the Bay: Annapolis, 

MD, 14 p.

Chesapeake Bay Program, 2006, Watershed model progress scenario results for 2005: 

Available on line at: http://www.chesapeakebay.net/tribtools.htm. 

Chesney, E.J., and Baltz, D.M., 2001, The effects of hypoxia on the northern Gulf of 

Mexico coastal ecosystem—A fisheries perspective, in Rabalais, N.N., and Turner, 

R.E., eds., Coastal hypoxia—Consequences for living resources and ecosystems: 

Washington, D.C., Coastal and Estuarine Studies 58, American Geophysical Union, 

p. 321–354. 

Childs, C.R., Rabalais, N.N., Turner, R.E., and Proctor, L.M., 2002, Sediment 

denitrification in the Gulf of Mexico zone of hypoxia: Marine Ecology Progress 

Series, v. 240, p. 285–290.

Childs, C.R., Rabalais, N.N., Turner, R.E., and Proctor, L.M., 2003, Erratum—Sediment 

denitrification in the Gulf of Mexico zone of hypoxia: Marine Ecology Progress 

Series, v. 247, p. 310.

Chow, T.L., Rees, H.W., and Daigle, J.L., 1999, Effectiveness of terraces/grassed 

waterway systems for soil and water conservation—A field evaluation: Journal of 

Soil and Water conservation, v. 54, no. 3, p. 577–583. 
233

Chua, T.T., Bronson, K.F., Booker, J.D., Keeling, J.W., Mosier, A.R., Bordovsky, J.P., 

Lascano, R.J., Green, C.J., and Segarra, E., 2003, In-season nitrogen status sensing 

in irrigated cotton— I. Yields and nitrogen-15 recovery: Soil Science Society of 

America Journal, v. 67, p. 1428-1438. 

Chung, S.W., Gassman, P.W., Gu, R., and Kanwar, R.S., 2002, Evaluation of EPIC for 

assessing tile flow and nitrogen losses for alternative agricultural management 

systems: Transactions of the American Society of Agricultural Engineers, v. 45, 

no. 4, p. 1135–1146. 

Claassen, R., 2000, Agricultural resources and environmental indicators—Compliance 

provisions for soil and wetland conservation: U.S. Department of Agriculture, 

Economic Research Service, Agricultural Resources and Environmental Indicators, 

chap. 6.3, 20 p.. 

Claassen, R., Breneman, V., Bucholtz, S., Cattaneo, A., Johansson, R., and Morehart, M., 

2004, Environmental compliance in U.S. agricultural policy—Past performance and 

future potential: U.S. Department of Agriculture, Economic Research Service, 

Agricultural Economic Report No. 832, p. 52. 

Clapp, C.E., Allmaras, R.R., Layese, M.F., Linden, D.R., and Dowdy, R.H., 2000, Soil 

organic carbon and 13C abundance as related to tillage, crop residue, and nitrogen 

fertilization under continuous corn management in Minnesota: Soil & Tillage 

Research, v. 55, no. 3, p. 127–142. 

Clark, C., and Russell, C., 2005, Public information provision as a tool of environmental 

policy?, in Krarup, S., and Russell, C., eds., Environment, information, and 

consumer behaviour, chap. 6: Cheltenham, United Kingdom, Edward Elgar 

Publishing. 

Cogeneration Technologies, 2006, Clean coal technology & the President’s clean coal 

power initiative: Available on line at 

http://www.cogeneration.net/IntegratedGasificationCombinedCycle.htm. 

Cohn, T.A., Caulder, D.L., Gilroy, E.J., Zynjuk, L.D., and Sommers, R.M., 1992, The 

validity of a simple statistical model for estimating fluvial constituent loads: An 

empirical study involving nutrient loads entering Chesapeake Bay: Water Resources 

Research, v. 28, p. 2352-2363. 

Collie, J.S., Richardson, K., and Steele, J.H., 2004, Regime shifts—Can ecological theory 

illuminate the mechanisms?: Progress in Oceanography, v. 60, p. 281–302. 

Conley, D.J., 2002, Terrestrial ecosystems and the global biogeochemical silica cycle: 

Global Biogeochemical Cycles, v. 16, p. 1121, doi:10.1029/2002GV001894. 

Conley, D. J., Carstensen, J., Ærtebjerg, G., Christensen, P.B., Dalsgaard, T., Hansen, 

J.L.S., and Josefson, A.B., 2007 Long-term changes and impacts of hypoxia in 

Danish coastal waters: Ecological Applications, Supplement, v. 17, no. 5, p. S165-

S184.

Conley, D.J., Humborg, C., Rahm, L., Savchuk, O.P., and Wulff, F., 2002a, Hypoxia in 

the Baltic Sea and basin-scale changes in phosphorus biogeochemistry: 

Environmental Science and Technology, v. 36, p. 5315–5320.
234

Conley, D.J., Markager, S., Andersen, J., Ellermann, T., and Svendsen, L.M., 2002b, 

Coastal eutrophication and the Danish National Aquatic Monitoring and Assessment 

Program: Estuaries, v. 25, p. 706–719.

Cooke, R.A., Sands, G.R., and Brown, L.C., in press, Drainage water management: a 

practice for reducing nitrate loads from subsurface drainage systems, in Proceedings 

of Gulf Hypoxia and Local Water Quality Concerns Workshop, Iowa State 

University, Ames, IA, September 26–28, 2005. (in press) 

Corbett, D.R., McKee, B.A., and Allison, M.A., 2006, Nature of decadal-scale sediment 

accumulation in the Mississippi river deltaic region: Continental Shelf Research, 

v. 26, p. 2125–2140.

Corbett, D.R., McKee, B.A., and Duncan, D., 2004, An evaluation of mobile mud 

dynamics in the Mississippi River deltaic region: Marine Geology, v. 209, p. 91–112. 

CPRA, 2007, Integrated ecosystem restoration and hurricane protection: Louisiana’s 

comprehensive master plan for a sustainable coast: Coastal Protection and 

Restoration Authority (CPRA) of Louisiana, Office of the Governor (Louisiana), 140 

p. Available on line at: http://www.lacpra.org/assets/docs/cprafinalreport5-2-07.pdf. 

Craig, J.K., and Crowder, L.B., 2005, Hypoxia-induced habitat shifts and energetic 

consequences in Atlantic croaker and brown shrimp on the Gulf of Mexico shelf: 

Marine Ecology Progressive Series, v. 294, p. 79-94.

Craig, J.K., Crowder, L.B., and Henwood, T.L., 2005, Spatial distribution of brown 

shrimp (Farfantepenaeus aztecus) on the northwestern Gulf of Mexico shelf—

Effects of abundance and hypoxia: Canadian Journal of Fisheries and Aquatic 

Sciences, v. 62, p. 1295-1308. 

Craig, J.K., Gray, C.D., McDaniel, C.M., Henwood, T.L., and Hanifen, J.G., 2001, 

Ecological effects of hypoxia on fish, sea turtles, and marine mammals in the 

northwestern Gulf of Mexico, in Rabalais, N.N., and Turner, R.E., eds., Coastal 

hypoxia—Consequences for living resources and ecosystems: Washington, D.C., 

Coastal and Estuarine Studies 58, American Geophysical Union, p. 269–291. 

Crespi, J., and Marette, S., 2005, Eco-labeling economics—Is public involvement 

necessary?, in Krarup, S., and Russell, C.S., eds., Environment, information, and 

consumer behaviour, chap. 5: Northhampton, Edward Elgar Publishing. 

Crumpton, W.G., Kovacic, D., Hey, D., and Kostel, J., in press, Potential of wetlands to 

reduce agricultural nutrient export to water resources in the corn belt, in Proceedings 

of Gulf Hypoxia and Local Water Quality Concerns Workshop, Iowa State 

University, Ames, IA, September 2005. (in press) 

Crumpton, W.G., Stenback, G.A., Miller, B.A., and Helmers, M.J., 2006, Potential 

benefits of wetland filters for tile drainage systems—Impact on nitrate loads to 

Mississippi river subbasins: U.S. Department of Agriculture, CSREES Project 

Completion Report, (release pending). 

Czapar, G.G., Payne, J., and Tate, J., 2007, An educational program on the proper timing 

of fall-applied nitrogen fertilizer: On-line Crop Management, doi: 10.194/CM-2007-

1510-01-RS. 
235

D’Sa, E.J., and Miller, R.L., 2003, Bio-optical properties in waters influenced by the 

Mississippi River during low flow conditions: Remote Sensing of Environment, 

v. 84, p. 538–549. 

Dagg, M., Benner, R., Lohrenz, S., and Lowrence, D., 2004, Transformation of dissolved 

and particulate materials on continental shelves influenced by large rivers—Plume 

processes: Continental Shelf Research, v. 24, p. 833–858.

Dagg, M.J., 1995, Copepod grazing and the fate of phytoplankton in the northern Gulf of 

Mexico: Continental Shelf Research, v. 15, nos. 11–12, p. 1303–1317.

Dagg, M.J., Ammerman, J.V., Amon, R., Gardner, W., Green, R., and Lohrenz, S., in 

press, Water column processes influencing hypoxia in the northern Gulf of Mexico: 

Estuaries and Coasts, v. 30, no. 4, p. __-__, in press. Accepted for publication notice 

last accessed August 16, 2007 at: 

http://estuariesandcoasts.org/contents/upcoming.html. (Draft 46 page manuscript 

available online at: http://www.epa.gov/msbasin/taskforce/pdf/session4dagg.pdf).

Dagg, M.J., and Breed, G.A., 2003, Biological effects of Mississippi River nitrogen on 

the northern Gulf of Mexico—A review and synthesis: Journal of Marine Systems, 

v. 43, p. 133–152.

Dagg, M.J., and Brown, S.L., 2005, The potential contribution of fecal pellets from the 

larvacean Oikopleura dioica to vertical flux of carbon in a river dominated coastal 

margin, in Gorsky, G., Youngbluth, M.J., and Deibel, D., eds., Response of marine 

ecosystems to global change: Ecological Impact of Appendicularians, Gordon and 

Breach, p. 293–307. 

Dahl, T.E., 1990, Wetlands losses in the United States—1780s to 1980s: Washington, 

D.C., U.S. Department of Interior, Fish and Wildlife Service, 21 p. 

Dale, V.H., Brown, S., Haeuber, R.A., Hobbs, N.T., Huntly, N., Naiman, R.J., Riebsame, 

W.E., Turner, M.G., and Valone, T.J., 2000, Ecological principles and guidelines for 

managing the use of land: Ecological Applications, v. 10, p. 639–670. 

Dalsgaard, T., Canfield, D.E., Petersen, J., Thamdrup, B., and Acuha-Gonzalez, J., 2003, 

N2 projection by the annamox reaction in the anoxic water column of Golfo Dulce, 

Costa Rica: Nature, v. 422, p. 606–608.

Dalzell, B.J., Gowda, P.H., and Mulla, D.J., 2004, Modeling sediment and phosphorus 

losses in an agricultural watershed to meet TMDLs: Journal of American Water 

Resources Association, v. 40, p. 533–543. 

Dampney, P.M.R., Lord, E.I., and Chambers, B.J., 2000, Development of advice for 

farmers and advisors: Soil Use and Management, v. 16, p. 162–166. 

Darrow, B.P., Walsh, J.J., Vargo, G.A., Masserini, R.T., Jr., Fanninga, K.A., and Zhang, 

J.-Z., 2003, A simulation study of the growth of benthic microalgae following the 

decline of a surface phytoplankton bloom: Continental Shelf Research, v. 23, 

p. 1265–1283.
236

David, M.B., and Gentry, L.E., 2000, Anthropogenic inputs of nitrogen and phosphorus 

and riverine export for Illinois, USA: Journal of Environmental Quality, v. 29, 

p. 494–508. 

David, M.B., Gentry, L.E., Kovacic, D.A., and Smith, K.M., 1997, Nitrogen balance in 

and export from an agricultural watershed: Journal of Environmental Quality, v. 26, 

p. 1038–1048. 

David, M.B., McIsaac, G.F., Royer, T.V., Darmody, R.G., and Gentry, L.E., 2001, 

Estimated historical and current nitrogen balances for Illinois: The Scientific World, 

v. 1, p. 597–604. 

David, M.B., Wall, L.G., Royer, T.V., and Tank, J.L., 2006, Denitrification and the 

nitrogen budget of a reservoir in an agricultural landscape: Ecological Applications, 

v. 16, p. 2177-2190. 

Davis, C.B., Baker, J.L., van der Valk, A.G., and Beer, C.E., 1981, Prairie pothole 

marshes as traps for nitrogen and phosphorus in agricultural runoff, in Richardson, 

B., ed., Proceedings of the Midwestern Conference on Wetland Values and 

Management, St. Paul, MN, 17–19 June 1981: Navarre, MN, Fresh Water Society, 

p. 153–163. 

Davis, J.G., Kitchen, N.R., Sudduth, K.A., and Drummond, S.T., 1997, Using 

electromagnetic induction to characterize soils: Potash & Phosphate Institute, Better 

Crops, v. 81, no. 4, p. 6–8. 

Day, J.W., Jr., Yanez Arancibia, A., Mitsch, W.J., Lara-Dominquez, A.L., Day, J.N., Ko, 

J-Y., Lane, R., and Lindsey, J., 2003, Using ecotechnology to address water quality 

and wetland habitat loss problems in the Mississippi basin—A hierarchical approach: 

Biotechnology Advances, v. 22, p. 135–159. 

Denbaly, M., and Vrooman, H., 1993, Dynamic fertilizer nutrient demands for corn: A 

co-integrated and error-correcting system: American Journal of Agricultural 

Economics, v. 75, p. 203–209. 

Deutsch, C., Sarmiento, J.L., Sigman, D.M., Gruber, N., and Dunne, J.P., 2007, Spatial 

coupling of nitrogen inputs and losses in the ocean: Nature, v. 445, p. 163–167, 

doi:10.1038/nature05392. 

Devlin, D., Dhuyvetter, K., McVay, K., Kastens, T., Rice, C., Janssen, K., and Pierznski, 

G., 2003, Water quality best management practices, effectiveness, and cost for 

reducing contaminant losses from cropland: Kansas State University Agricultural 

Experiment Station and Cooperative Extension Service MF-2572, 4 p. 

Diaz, R.J., 2001, Overview of hypoxia around the world: Journal of Environmental 

Quality, v. 30, no. 2, p. 275–281.

Diaz, R.J., Nestlerode, J., and Diaz, M.L., 2003, A global perspective on the effects of 

eutrophication and hypoxia on aquatic biota, in Proceedings of the Seventh 

International Symposium, Fish Physiology, Toxicology, and Water Quality, May 12–

15, 2003: Tallinn, Estonia.
237

Diaz, R.J., and Rosenberg, R., 1995, Marine benthic hypoxia—A review of its ecological 

effects and the behavioral responses of benthic macrofauna: Oceanography & 

Marine Biology, An Annual Review, v. 33, p. 245–303.

Diaz, R.J., and Solow, A., 1999, Ecological and economic consequences of hypoxia—

Topic 2, Gulf of Mexico hypoxia assessment: Silver Springs, MD, National Ocean 

and Atmospheric Administration Coastal Ocean Program Decision Analysis Series, 

86 p., available online at: http://oceanservice.noaa.gov/products/hypox_t2final.pdf.

DiMarco, S., Hetland, R., Howden, S., Murray, S., Walker, N., and Wiseman, W., 2007, 

Influence of physical oceanographic processes on the distribution and extent of the 

hypoxic zone: (In review) 

DiMarco, S.F., Walker, N., Wiseman, W.J., Jr., Murray, S.P., and Howden, S.D., 2006, 

Physical processes of the northern Gulf of Mexico and their influence on hypoxia of 

the Texas-Louisiana shelf: 

Dinnel, S.P., and Wiseman, W.J., 1986, Fresh-water on the Louisiana and Texas shelf: 

Continental Shelf Research, v. 6, p. 765–784. 

Dinnes, D.L., 2004, Assessments of practices to reduce nitrogen and phosphorus 

nonpoint source pollution of Iowa’s surface waters: Ames, IA, U.S. Department of 

Agriculture, Agricultural Research Service, National Soil Tilth Laboratory, 376 p. 

Djodjic, F., Ulen, B., and Bergstrom, L., 2000, Temporal and spatial variations of 

phosphorus losses and drainage in a structured clay soil: Water Research, v. 34, 

p. 1687–1695. 

Dobermann, A., Walters, D.T., and Adviento-Borbe, M.A.A., 2007, Global warming 

potential of high-yielding continuous corn and corn-soybean systems: Better Crops, 

v. 91, no. 3, p. 16–19. 

Dodds, W.K., Martí, E., Tank, J.L., Pontius, J., Hamilton, S.K., Grimm, N.B., Bowden, 

W.B., McDowell, W.H., Peterson, B.J., Valett, H.M., Webster, J.R., and Gregory, S., 

2004, Carbon and nitrogen stoichiometry and nitrogen cycling rates in streams: 

Oceologia, v. 140, p. 458-467. 

Doering, O., Diaz-Hermelo, F., Howard, C., Heimlich, R., Hitzhusen, F., Kazmierczak, 

R., Lee, J., Libby, L., Milon, W., Prato, T., and Ribaudo, M., 1999, Evaluation of the 

economic costs and benefits of methods for reducing nutrient loads to the Gulf of 

Mexico: Topic 6 report for the integrated assessment of hypoxia in the Gulf of 

Mexico: Silver Springs, MD, National Oceanic and Atmospheric Administration 

Coastal Ocean Program, Decision Analysis Series No. 20, 137 p. Available on line 

at: http://oceanservice.noaa.gov/products/hypox_t6final.pdf. 

Doering, O.C., 2002, Economic linkages driving the potential response to nitrogen overenrichment: Estuaries, v. 25, no. 4B, p. 809– 818. 

Donner, S.D., 2006, Surf or turf—A shift from feed to food cultivation could reduce 

nutrient flux to the Gulf of Mexico: Global Environmental Change, v. 17, no. 1, p. 

105-113. 
238

Donner, S.D., Coe, M.T., Lenters, J.D., Twine, T.E., and Foley, J.A., 2002, Modeling the 

impact of hydrological changes on nitrate transport in the Mississippi River Basin 

from 1955 to 1994: Global Biogeochemical Cycles, v. 16, 

doi:10.1029/2001GB001396. 

Donner, S.D., and Kucharik, C.J., 2003a, Evaluating the impacts of land management and 

climate variability on crop production and nitrogen export across the Upper 

Mississippi Basin: Global Biogeochemical Cycles, v. 17, doi:10.1028/2001GB1808. 

Donner, S.D., and Kucharik, C.J., 2003b, The distribution of the primary crops in the 

U.S. since 1950 and the relationship to river nutrient levels: Global Biogeochemical 

Cycles, v. 17, doi:10.1029/2001GB1808. 

Donner, S.D., Kucharik, C.J., and Foley, J.A., 2004, Impact of changing land use 

practices on nitrate export by the Mississippi River: Global Biogeochemical Cycles, 

v.18, no. GB1028. 

Donner, S.D., and Scavia, D., 2007, How climate controls the flux of nitrogen by the 

Mississippi River and the development of hypoxia in the Gulf of Mexico: Limnology 

and Oceanography, v. 52, no. 2, p. 856–861.

Dortch, Q., Rabalais, N.N., Turner, R.E., and Qureshi, N.A., 2001, Impacts of changing 

Si/N ratios and phytoplankton species composition, in Rabalais, N.N., and Turner, 

R.E., eds., Coastal hypoxia—Consequences for living resources and ecosystems: 

Washington, D.C., Coastal and Estuarine Studies 58, American Geophysical Union, 

p. 37–48.

Dortch, Q., and Whitledge, T.E., 1992, Does nitrogen or silicon limit phytoplankton 

production in the Mississippi River plume and nearby regions?: Continental Shelf 

Research, v. 12, p. 1293–1309.

Dosskey, M.G., 2001, Toward quantifying water pollution abatement in response to 

installing buffers on crop land: Environmental Management, v. 28, no. 5, p. 577–

598. 

Dou, Z., Lanyon, L.E., Ferguson, J.D., Kohn, R.A., Boston, R.C., and Chalupa, W., 1998, 

An integrated approach to managing nitrogen on dairy farms—Evaluating farm 

performance using dairy nitrogen planner: Agronomy Journal, v. 90, p. 573–581. 

Du, B., Saleh, A., Jaynes, D.B., and Arnold, J.G., 2006, Evaluation of SWAT in 

simulating nitrate nitrogen and atrazine fates in a watershed with tiles and potholes: 

Transactions of the American Society of Agricultural and Biological Engineers, 

v. 49, no. 4, p. 949–959. 

Duan, S., and Bianchi, T.S., 2006, Seasonal changes in the abundance and composition of 

plant pigments in particulate organic carbon in the lower Mississippi and Pearl 

Rivers: Estuaries and Coasts, v. 29, no. 3, p. 427–442. 

Duan, S., Bianchi, T.S., and Sampere, T.P., 2007, Temporal variability in the 

composition and abundance of terrestrially-derived dissolved organic matter in the 

lower Mississippi and Pearl Rivers: Marine Chemistry, v. 103, p. 172–184.
239

Edwards, D.R., and Daniel, T.C., 1993a, Drying-interval effects on runoff from fescue 

plots receiving swine manure: Transactions of the American Society of Agricultural 

Engineers, v. 36, p. 1673–1678. 

Edwards, D.R., and Daniel, T.C., 1993b, Runoff quality impacts of swine manure applied 

to fescue plots: Transactions of the American Society of Agricultural Engineers, 

v. 36, p. 81–80. 

Eldridge, P.M., and Morse, J.W., in press, Origins and temporal scales of hypoxia on the 

Louisiana shelf—Importance of benthic and sub-pycnocline water metabolism: (in 

press). 

Elobeid, A., Tokgoz, S., Hayes, D.J., Babcock, B.A., and Hart, C.E., 2006, The long-run 

impact of corn-based ethanol on the grain, oilseed and livestock sectors: A 

preliminary assessment: Ames, IA, Center of Agriculture and Rural Development 

Briefing Paper 06-BP 49. 

Engelhaupt, E., and Bianchi, T.S., 2001, Sources and composition of high-molecularweight dissolved organic carbon in a southern Louisiana tidal stream (Bayou 

Trepagnier): Limnology and Oceanography, v. 46, p. 917–926.

English, B.C., De la Torre Ugarte, D.G., Jensen, K., Hellwinckel, C., Menard, J., Wilson, 

B., Roberts, R., and Walsh, M., 2006, 25% renewable energy for the Unites States by 

2025: Agricultural and economic impacts: Knoxville, TN, University of Tennessee 

Technical Report. 

Etter, P.C., Howard, M.K., and Cochrane, J.D., 2004, Heat and freshwater budgets of the 

Texas-Louisiana shelf: Journal of Geophysical Research, v. 109, C02024, 

doi:10.1029/2003JC001820. 

Farrell, A.E., Plevin, R.J., Turner, B.T., Jones, A.D., O’Hare, M., and Kammen, D.M., 

2006, Ethanol can contribute to energy and environmental goals: Science, v. 311, no. 

5760, p. 506-508. 

Feather, P., Hellerstein, D., and Hansen, L., 1999, Economic valuation of environmental 

benefits and the targeting of conservation programs—The case of the CRP: U.S. 

Department of Agriculture, Economic Research Service, Agricultural Economics 

Report No. 778, 64 p. 

Feng, H., Kurkalova, L.A., Kling, C.L., and Gassman, P.W., 2005, Economic and 

environmental co-benefits of carbon sequestration in agricultural soils—Retiring 

agricultural land in the upper Mississippi River basin: Iowa State University Center 

for Agricultural and rural Development (CARD) Publication 05-WP384, 23 p., 

available online at: 

http://www.card.iastate.edu/publications/DBS/PDFFiles/05wp384.pdf.

Feng, H., Kurkalova, L.A., Kling, C.L., and Gassman, P.W., 2006, Environmental 

conservation in agriculture: Land retirement vs. changing practices on working land: 

Journal of Environmental Economics and Management, v. 52, p. 600-614. 

Ferguson, R.B., Lark, R.M., and Slater, G.P., 2003, Approaches to management zone 

definition for use of nitrification inhibitors: Soil Science Society of America Journal, 

v. 67, p. 937-947. 
240

Fiener, P., and Auerswald, K., 2003, Effectiveness of grassed waterways in reducing 

runoff and sediment delivery from agricultural watersheds: Journal of Environmental 

Quality, v. 32, no. 3, p. 927–936. 

Fisher, T.R., and Gustafson, A.B., 2004, Progress Report—Aug. 1990–Dec. 2003, 

Nutrient-addition bioassays in Chesapeake Bay to assess resources limiting 

phytoplankton growth: Annapolis, MD, Maryland Department of Natural Resources, 

50 p.

Fisher, T.R., Peele, E.R., Ammerman, J.W., and Harding, L.W., 1992, Nutrient limitation 

of phytoplankton in Chesapeake Bay: Marine Ecology Progress Series, v. 82, p. 51–

63. 

Forman, R.T.T., 1995, Land mosaics—The ecology of landscapes and regions: 

Cambridge, England, Cambridge University Press, 652 p. 

Fox, L., Sager, S.L., and Wofsy, S.C., 1985, Factors controlling the concentrations of 

soluble phosphorus in the Mississippi estuary: Limnology & Oceanography, v. 30, 

p. 826–832.

Frankenberger, J., Kladivko, E., Sands, G., Jaynes, D.B., Fausey, N.R., Helmers, M., 

Cooke, R., Strock, J., Nelson, K., and Brown, L., 2006, Drainage water management 

for the Midwest: Purdue Extension: Knowledge to Go, WQ-44, 8 p., available online 

at: http://www.ces.purdue.edu/extmedia/WQ/WQ-44.pdf. 

Froelich, P.N., 1988, Kinetic control of dissolved phosphate in natural rivers and 

estuaries—A primer on the phosphate buffer mechanism: Limnology & 

Oceanography, v. 33, p. 649–668.

Galler, J.J., Bianchi, T.S., Allison, M.A., Campanella, R., and Wysocki, L., 2003, 

Sources of aged terrestrial organic carbon to the Gulf of Mexico from relict strata in 

the Mississippi River: EOS, Transactions of the American Geophysical Union, v. 84, 

p. 469–476.

Gardner, G., 1998, Recycling organic wastes, in Brown, L., Flavin, C., and French, H., 

eds., State of the world: New York, NY, W.W. Norton, p. 96–112. 

Gardner, W.S., Benner, R., Chin-Leo, G., Cotner, J.B., Eadie, B.J., Cavaletto, J.F., 

Lansing, M.B., 1994, Mineralization of organic material and bacterial dynamics in 

Mississippi River plume water: Estuaries, v. 17, no. 4, p. 816–828.

Garnier, J., Leporcq, B., Sanches, N., and Philippon, X., 1999, Biogeochemical mass 

balances (C, N, P, Si) in three large reservoirs of the Seine Basin (France): 

Biogeochemistry, v. 47, p. 119–146. 

Garnier, J., Sferratore, A., Meybeck, M., Billen, G., and Durr, H., 2006, Modeling silica 

transfer process in watersheds, in Ittekot, V., Unger, D., Humborg, C., and An, N.T., 

eds., The silicon cycle—Human perturbations and impacts on aquatic systems: 

Washington, D.C., Island Press, p. 139–162. 

Gassman, P.W., Reyes, M.R., Green, C.H., and Arnold, J.G., 2007, The soil and water 

assessment tool—Historical development applications, and future research 
241

directions: Transactions of the American Society of Agricultural and Biological 

Engineers, v.50, no. 4, p. 1211-1240. 

Gehl, R.J., Schmidt, J.P., Maddux, L.D., and Gordon, W.B., 2005, Corn yield response to 

nitrogen rate and timing in sandy irrigated soils: Agronomy Journal, v. 97, p. 1230–

1238. 

Gentry, L.E., Below, F.E., David, M.B., and Bergerou, J.A., 2001, Source of the soybean 

N credit in maize production: Plant and Soil, v. 236, p. 175–184. 

Gentry, L.E., David, M.B., Royer, T.V., Mitchell, C.A., and Starks, K.M., 2007, 

Phosphorus transport pathways to streams in tile-drained agricultural watersheds: 

Journal of Environmental Quality, v. 36, p. 408–415. 

Gilroy, E.J., Hirsh, R.M., and Cohn, T.A., 1990, Mean square error of regression-based 

constituent transport estimates: Water Resources Research, v. 26, p. 2069-2077. 

Gitau, M.W., Gburek, W.J., and Jarrett, A.R., 2005, A tool for estimating best 

management practice effectiveness for phosphorus pollution control: Journal of Soil 

and Water Conservation, v. 60, no. 1, p. 1–10. 

Gollehon, N., Caswell, M., Ribaudo, M., Kellogg, R., Lander, C., and Letson, D., 2001, 

Confined animal production and manure nutrients: U.S. Department of Agriculture, 

Economic Research Service, Resource Economics Division, Agriculture Information 

Bulletin No. 771. 

Goni, M.A., Ruttenberg, K.C., and Eglinton, T.I., 1998, A reassessment of the sources 

and importance of land-derived organic matter in surface sediments from the Gulf of 

Mexico: Geochimica et Cosmochimica Acta, v. 62, no. 18, p. 3055-3075. 

Goolsby, D.A., Battaglin, W.A., Aulenbach, B.T., and Hooper, R.P., 2001, Nitrogen 

input to the Gulf of Mexico: Journal of Environmental Quality, v. 30, no. 2, p. 329–

336.

Goolsby, D.A., Battaglin, W.A., Lawrence, G.B., Artz, R.S., Aulenbach, B.T., Hooper, 

R.P., Keeney, D.R., and Stensland, G.S., 1999, Flux and sources of nutrients in the 

Mississippi–Atchafalaya River basin—Topic 3, Report for the integrated assessment 

of hypoxia in the Gulf of Mexico: National Oceanic and Atmospheric Administration 

Coastal Ocean Program Decision Analysis Series 17, 128 p. 

Gordon, E.S., Goni, M.A., Roberts, Q.N., Kineke, G.C., and Allison, M.A., 2001, 

Organic matter distribution and accumulation on the inner Louisiana shelf west of 

the Atchafalaya River: Continental Shelf Research, v. 21, p. 1691–1721.

Gowda, P.H., Dalzell, B.J., and Mulla, D.A., 2007, Model based nitrate TMDLs for two 

agricultural watersheds of southeastern Minnesota: Journal of the American Water 

Resources Association, v. 43, no. 1, p. 254–263. 

Gowda, P.H., and Mulla, D.J., 2006, Modeling alternative agricultural management 

practices for High Island Creek watershed in South-Central Minnesota: Journal of 

Environmental Hydrology, v. 14, no. 13, p. 1–15. 

Graco, M., Farías, L., Molina, V., Gutiérrez, D., and Nielsen, L.P., 2001, Massive 

developments of microbial mats following phytoplankton blooms in a naturally 
242

eutrophic bay—Implications for nitrogen cycling: Limnology & Oceanography, 

v. 46, p. 821–832.

Graff, C.D., Sadeghi, A.M., Lowrance, R.R., and Williams, R.G., 2005, Quantifying the 

sensitivity of the riparian ecosystem management model (REMM) to changes in 

climate and buffer characteristics common to conservation practices: Transactions of 

the American Society of Agricultural Engineers, v. 48, no. 4, p. 1377–1387. 

Grandy, A.S., Loecke, T.D., Parr, S., and Robertson, G.P., 2006, Long-term trends in 

nitrous oxide emissions, soil nitrogen, and crop yields of till and no-till cropping 

systems: Journal of Environmental Quality, v. 35, p. 1487-1495. 

Graham, R.L., Nelson, R., Sheehan, J., Perlack, R.D., and Wright, L.L., 2007, Current 

and potential U.S. corn stover supplies: Agronomy Journal, v. 99, p. 1-11. 

Green, C.H., Tomer, M.D., DiLuzio, M., and Arnold, J.G., 2006a, Hydrologic evaluation 

of the soil and water assessment tool for a large tile-drained watershed in Iowa: 

Transactions of the American Society of Agricultural and Biological Engineers, 

v. 49, no. 2, p. 413–422. 

Green, R.E., Bianchi, T.S., Dagg, M.J., Walker, N.D., and Breed, G.A., 2006b, An 

organic carbon budget for the Mississippi River turbidity plume and plume 

contributions to air-sea CO2 fluxes and bottom water hypoxia: Estuaries, v. 29, no. 4, 

p. 579–597.

Green, R.E., Breed, G.A., Dagg, M.J., and Lohrenz, S.E., in press, Modeling the response 

of primary production and sedimentation to variable nitrate loading in the 

Mississippi River plume: (In review). 

Green, W.R., and Haggard, B.E., 2001, Phosphorus and nitrogen concentrations and 

loads at Illinois River south of Siloam Springs, Arkansas, 1997–1999: U.S. 

Geological Survey Water Resources Investigations Report 01-4217, 12 p., available 

online at http://pubs.er.usgs.gov/usgspubs/wri/wri014217

Greenhalgh, S., and Sauer, A., 2003, Awakening the dead zone: An investment for 

agriculture, water quality, and climate changes: World Resources Institute, WRI 

Issue Brief, February, 24 p. 

Gregg, M.C., 1999, Uncertainties and limitations in measuring ε and χT: Journal of 

Atmospheric and Oceanic Technology, v. 16, p. 1483–1490. 

Grizzetti, B., Bouraoui, F., and De Marsily, G., 2005, Modeling nitrogen pressure in river 

basins—A comparison between a statistical approach and the physically-based 

SWAT model: Physics and Chemistry of the Earth, Parts A/B/C, v. 30, nos. 8–10, 

p. 508–517. 

Groffman, P., Altabet, M.A., Böhlke, J.K., Butterbach-Bahl, K., David, M.B., Firestone, 

M.K., Giblin, A.E., Kana, T.M., Nielsen, L.P., and Voytek, M.A., 2006, Methods for 

measuring denitrification—Diverse approaches to a difficult problem: Ecological 

Applications, v. 16, p. 2091-2122, available online at: 

https://darchive.mblwhoilibrary.org/bitstream/1912/1425/1/Groffman%20et%20al.%

20%282006%29%20-%20denitrification%20methods%20paper.pdf. 
243

Hagy, J.D., Boynton, W.R., Keefe, C.W., and Wood, K.V., 2004, Hypoxia in Chesapeake 

Bay, 1950–2001—Long-term change in relation to nutrient loading and river flow: 

Estuaries, v. 27, p. 634–658.

Harding, L.W., Jr., 1994, Long-term trends in the distribution of phytoplankton in 

Chesapeake Bay—Roles of light, nutrients, and streamflow: Marine Ecology 

Progress Series, v. 104, p. 267–291.

Harding, L.W., Jr., Mallonee, M.E., and Perry, E.S., 2002, Toward a predictive 

understanding of primary productivity in a temperate, partially stratified estuary: 

Estuarine Coastal & Shelf Science, v. 55, p. 437–463.

Haufler, J., ed., 2005, Fish and wildlife benefits of Farm Bill Conservation Programs—

2000–2005 update: The Wildlife Society, Technical Review 05-2. 

Haygarth, P.M., Chapman, P.J., Jarvis, S.C., and Smith, R.V., 1998, Phosphorus budgets 

for two contrasting grassland farming systems in the UK: Soil Use and Management, 

v. 14, p. 160–167. 

Heckrath, G., Brookes, P.C., Paulton, P.R., and Goulding, K.W.T., 1995, Phosphorus 

leaching from soils containing different phosphorus concentrations in the Broadbalk 

experiment: Journal of Environmental Quality, v. 24, p. 904–910. 

Helmers, M., 2007, Iowa State University agricultural drainage: Available at 

http://www3.abe.iastate.edu/agdrainage/. 

Helsel, D. R., and Hirsch, R.M., 2002, Statistical methods in water resources: U.S. 

Geological Survey Techniques of Water-Resources Investigations, Book 4, Chapter 

A3, available at http://water.usgs.gov/pubs/twri/twri4a3/. 

Helton, A.M., 2006, An inter-biome comparison of stream network nitrate dynamics: 

Athens, GA, University of Georgia, M.S. thesis, 143 p. 

Hendon, L.A., Carlson, E.A., Manning, S., and Brouwer, M., 2006, Cross-talk between 

pyrene and hypoxia signaling pathways in embryonic Cyprinodon variegates: Paper 

presented at Hypoxia Effects on Living Resources in the Gulf of Mexico, September 

25–26, 2006: New Orleans, Louisiana, Tulane University, sponsored by National 

Oceanic and Atmospheric Administration Center for Sponsored Coastal Ocean 

Research. 

Herbert, R.A., 1999, Nitrogen cycling in coastal marine systems: FEMS Microbiology 

Reviews, v. 23, p. 563–590.

Hernandez, M.E., and Mitsch, W.J., 2006, Influence of hydrologic pulses, flooding 

frequency, and vegetation on nitrous oxide emissions from created riparian marshes: 

Wetlands, v. 26, no. 3, p. 862-877. 

Hernes, P.J., and Benner, R., 2003, Photochemical and microbial degradation of 

dissolved lignin phenols—Implications for the fate of terrigenous dissolved organic 

matter in marine environments: Journal of Geophysical Research—Oceans, 

v. 108(C9).
244

Hetland, R.D., and DiMarco, S.F., 2007, How does the character of oxygen demand 

control the structure of hypoxia on the Texas-Louisiana continental shelf?: Journal of 

Marine Systems, doi:10.1016/j.jmarsys.2007.03.002.

Hey, D.L., Kenimer, A.L., and Barrett, K.R., 1994, Water quality improvement by four 

experimental wetlands: Ecological Engineering, v. 3, p. 381–397. 

Hey, D.L., Montgomery, D.L., Urban, L.S., Prato, T., Zarwell, R., Forbes, A., Martell, 

M., Pollack, J., and Steele, Y., 2004, Flood damage reduction in the Upper 

Mississippi River Basin—An ecological alternative: The Wetlands Initiative, 

Chicago, IL, 44 p., available online at: http://www.wetlandsinitiative.org/images/UMRBFinalReport.pdf. 

Hoffman, B.S., Brouder, S.M., and Turco, R.F., 2004, Tile spacing impacts on Zea mays

L. yield and drainage water nitrate load: Ecological Engineering, v. 23, p. 251–267. 

Holland, E.F., Braswell, B.H., Sulzman, J., and Lamarque, J., 2005, Nitrogen deposition 

to the United States and Western Europe—Synthesis of observations and models: 

Ecological Applications, v. 15, p. 38–57. 

Holland, E.F., Dentener, B., Braswell, B.H., and Sulzman, J., 1999, Contemporary and 

pre-industrial global reactive nitrogen budgets: Biogeochemistry, v. 4, p. 7–43. 

Hong, N., White, J.G., Weisz, R., Crozier, C.R., Gumpertz, M.L., and Cassel, D.K., 2006, 

Remote sensing-informed variable-rate nitrogen management of wheat and corn—

Agronomic and groundwater outcomes: Agronomy Journal, v. 98, p. 327–338. 

Howarth, R.W., 1984, The ecological significance of sulfur in the energy dynamics of 

salt marsh and marine sediments: Biogeochemistry, v. 1, p. 5–27.

Howarth, R.W., 1998, An assessment of human influences on inputs of nitrogen to the 

estuaries and continental shelves of the North Atlantic Ocean: Nutrient Cycling in 

Agroecosystems, v. 52, p. 213-223. 

Howarth, R.W., 2006, Atmospheric deposition and nitrogen pollution in coastal marine 

ecosystems, in Visgilio, G.R., and Whitelaw, D.M., eds., Acid in the environment—

Lessons learned and future prospects: NY, Springer, p. 97–116. 

Howarth, R.W., Billen, G., Swaney, D., Townsend, A., Jarworski, N., Lajtha, K., 

Downing, J.A., Elmgren, R., Caraco, N., Jordan, T., Berendse, F., Freney, J., 

Kueyarov, V., Murdoch, P., and Zhao-liang, Zhu., 1996, Riverine inputs of nitrogen 

to the North Atlantic Ocean—Fluxes and human influences: Biogeochemistry, v. 35, 

p. 75–139. 

Howarth, R.W., Boyer, E.W., Pabich, W.J., and Galloway, J.N., 2002, Nitrogen use in the 

United States from 1961–2000 and potential future trends: Ambio, v. 31, p. 88–96. 

Howarth, R.W., Jensen, H.S., Marino, R., and Postma, H., 1995, Transport to and 

processing of P in near-shore and oceanic waters, in Tiessen, H., ed., Phosphorus in 

the global environment: Wiley, p. 323–345.

Howarth, R.W., and Marino, R., 2006, Nitrogen as the limiting nutrient for eutrophication 

in coastal marine ecosystems—Evolving views over 3 decades: Limnology & 

Oceanography, v. 51, p. 364–376. 
245

Howarth, R.W., Marino, R., Swaney, D.P., and Boyer, E.W., 2006a, Wastewater and 

watershed influences on primary productivity and oxygen dynamics in the lower 

Hudson River estuary, in Levinton, J.S., and Waldman, J.R., eds., The Hudson River 

Estuary: Cambridge University Press, p. 121–139. 

Howarth, R.W., Ramakrishna, K., Choi, E., Elmgren, R., Martinelli, L., Mendoza, A., 

Moomaw, W., Palm, C., Boy, R., Scholes, M., and Zhao-Liang, Zhu, 2005, Nutrient 

management, responses assessment, in Ecosystems and human well-being—Volume 

3, Policy responses, The millennium ecosystem assessment: Washington, D.C., 

Island Press, chap. 9, p. 295–311. 

Howarth, R.W., Swaney, D.P., Boyer, E.W., Marino, R., Jaworski, N. and Goodale, C., 

2006b, The influence of climate on average nitrogen export from large watersheds in 

the Northeastern United States: Biogeochemistry, v. 79, p 163–186. 

Hu, X., McIssac, G.F., David, M.B., and Louwers, C.A.L., 2007, Modeling riverine 

nitrate export from an east-central Illinois watershed using SWAT: Journal of 

Environmental Quality, v. 36, p. 996–1005. 

Humborg, C., Conley, D.J., Rahm, L., Wulff, F., Cociasu, A., and Ittekkot, V., 2000, 

Silicon retention in river basins—Far-reaching effects on biogeochemistry and 

aquatic food webs in coastal marine environments: Ambio, v. 29, p. 45–50. 

IATP, 2006, Staying home: How ethanol will change U.S. corn exports: Minneapolis, 

MN, Institute for Agriculture and Trade Policy, 26 p.

Inwood, S.E., Tank, J.L., and Bernot, M.J., 2005, Patterns of denitrification associated 

with land use in 9 mid-western headwater streams: Journal of the North American 

Benthological Society, v. 24, p. 227–245. 

Isik, M., and Khanna, M., 2002, Variable rate nitrogen application under uncertainty: 

Implications for profitability and nitrogen use: Journal of Agricultural and Resource 

Economics, v. 27, no. 1, p. 61-76. 

Isik, M., and Khanna, M., 2003, Stochastic technology, risk preferences and adoption of 

site-specific technologies: American Journal of Agricultural Economics, v. 85, no. 2, 

p. 305-317. 

James, E.E., 2005, Factors influencing the adoption and non-adoption of the conservation 

reserve enhancement program in the Cannonsville Watershed, New York: University 

Park, PA, The Pennsylvania State University, Department of Agricultural Economics 

and Rural Sociology, M.Sc. dissertation, 174 p. 

James, E.E., Kleinman, P.J.A., Veith, T.L., Stedman, R., and Sharpley, A.N., 2007, 

Phosphorus contributions from pastured dairy cattle to streams in the Cannonsville 

watershed: Journal of Soil and Water Conservation, v. 62, p. 40-47. 

Jarosz, E., and Murray, S.P., 2005, Velocity and transport characteristics of the 

Louisiana-Texas coastal current: Geophysical Monograph, American Geophysical 

Union, Circulation in the Gulf of Mexico Observations and Models, v. 161, p. 143–

156. 
246

Jaynes, D.B., Colvin, T.S., Karlen, D.L., Cambardella, C.A., and Meek, D.W., 2001, 

Nitrate loss in subsurface drainage as affected by nitrogen fertilizer rate: Journal of 

Environmental Quality, v. 30, p. 1305–1314. 

Jaynes, D.B., Dinnes, D.L., Meek, D.W., Karlen, D.L., Cambardella, C.A., and Colvin, 

T.S., 2004, Using the late spring nitrate test to reduce nitrate loss within a watershed: 

Journal of Environmental Quality, v. 33, p. 669–677. 

Jaynes, D.B., and James, D.E., 2007, The extent of farm drainage in the United States: 

Poster presentation at Soil and Water Conservation Society 2007 Annual 

Conference, July 21-25, 2007, Tampa, Florida, online at: 

http://www.ars.usda.gov/Research/Researchltrm?modecode=36-25-15-00. 

Jaynes, D.B., and Karlen, D.L., 2005, Sustaining soil resources while managing nutrients, 

in Proceedings of the Upper Mississippi River Sub-Basin Hypoxia Nutrient 

Committee Workshop, September 26–28, 2005, Ames, Iowa, p. 141–150, accessed 

June 28, 2007, at www.umrshnc.org/files/Hypwebversion.pdf.

Jha, M., Arnold, J.G., Gassman, P.W., Giorgi, F., and Gu, R.R., 2006, Climate change 

sensitivity assessment on Upper Mississippi River Basin streamflows using SWAT: 

Journal of American Water Resources Association, v. 42, no. 4, p. 997–1016. 

Jochem, F.J., McCarthy, M.J., and Gardner, W.S., 2004, Microbial ammonium recycling 

in the Mississippi River plume during the drought spring of 2000: Journal of 

Plankton Research, v. 26, p. 1265–1275. 

Joye, S.B., and Hollibaugh, J.T., 1995, Influence of sulfide inhibition on nitrification and 

nitrogen regeneration in sediments: Science, v. 270, p. 623–625. 

Justić, D., Bierman, V.J., Jr., Scavia, D., and Hetland, R., in press, Forecasting Gulf’s 

hypoxia: The next 50 years?: (in press).

Justić, D., Rabalais, N.N., and Turner, R.E., 1996, Effects of climate change on hypoxia 

in coastal waters—A doubled CO2 scenario for the northern Gulf of Mexico: 

Limnology & Oceanography, v. 41, p. 992–1003. 

Justić, D., Rabalais, N.N., and Turner, R.E., 2002, Modeling the impacts of decadal 

changes in riverine nutrient fluxes on coastal eutrophication near the Mississippi 

River delta: Ecological Modeling, v. 152, p. 33–46.

Justić, D., Rabalais, N.R., and Turner, R.E., 2003a, Simulated responses of the Gulf of 

Mexico hypoxia to variations in climate and anthropogenic nutrient loading: Journal 

of Marine Systems, v. 42, p. 115–126. 

Justić, D., Rabalais, N.N., Turner, R.E., and Dortch, Q., 1995, Changes in nutrient 

structure of river-dominated coastal waters—Stoichiometeric nutrient balance and its 

consequences: Estuarine, Coastal Shelf Science, v. 40, p. 339–356.

Justić, D., Rabalais, N.N., Turner, R.E., and Wiseman, W.J., Jr., 1993, Seasonal coupling 

between riverborne nutrients, net productivity and hypoxia: Marine Pollution 

Bulletin 26, p. 184–189. 
247

Justić, D., Turner, R.E., and Rabalais, N.N., 2003b, Climate influences on riverine nitrate 

flux—Implications for coastal marine eutrophication and hypoxia: Estuaries, v. 26, 

no. 1, p. 1–11.

Kantha, L.H., 2005, Barotropic tides in the Gulf of Mexico, in Sturges, W., and LugoFernandez, A., eds., Circulation in the Gulf of Mexico—Observations and models: 

American Geophysical Union, Geophysical Monograph, v. 161, p. 159–163. 

Kantha, L.H., and Clayson, C.A., 2000, Small scale processes in geophysical fluid flows: 

San Diego, CA, Academic Press, 668p. 

Kanwar, R.S., Colvin, T.S., and Karlin, D.L., 1997, Ridge, moldboard, chisel and no-till 

effects on tile water quality beneath two cropping systems: Journal of Production 

Agriculture, v. 10, p. 227-234. 

Karlen, D.L., Dinnes, D.L., Jaynes, D.B., Hurburgh, C.R., Cambardella, C.A., Colvin, 

T.S., and Rippke, G.R., 2005, Corn response to late-spring nitrogen management in 

the Walnut Creek watershed: Agronomy Journal, v. 97, p. 1054–1061. 

Karlen, D.L., Kramer, L.A., and Logsdon, S.D., 1998a, Field-scale nitrogen balances 

associated with long-term continuous corn production: Agronomy Journal, v. 90, 

p. 644–650. 

Karlen, D.L., Kumar, A., Kanwar, R.S., Cambardella, C.A., and Colvin, T.S., 1998b, 

Tillage system effects on 15 year carbon-based and simulated N budgets in a tiledrained Iowa field: Soil & Tillage Research, v. 48, p. 155–165. 

Karlson, K., Rosenberg, R., and Bonsdorff, E., 2002, Temporal and spatial large-scale 

effects of eutrophication and oxygen deficiency on benthic fauna in Scandinavian 

and Baltic waters—A review: Oceanography & Marine Biology, An Annual Review, 

v. 40, p. 427–489. 

Keene, W.C., Montag, J.A., Maben, J.A., Southwell, M., Leonard, J., Church, T.M., 

Moody, J.L., and Galloway, J.N., 2002, Organic nitrogen in precipitation over 

eastern North America: Atmospheric Environment, v. 36, p. 4529–4540. 

Kellogg, L.E., and Bridgham, S.D., 2003, Phosphorus retention and movement across an 

ombrotrophic-minerotrophic peatland gradient: Biogeochemistry, v. 63, p. 299–315. 

Kellogg, R.L., Lander, C.H., Moffitt, D.C., and Gollehon, N., 2000, Manure nutrients 

relative to the capacity of cropland and pastureland to assimilate nutrients—Spatial 

and temporal trends for the United States: U.S. Department of Agriculture, Natural 

Resources Conservation Service and Economic Research Service, Resource 

Assessment and Strategic Planning Working Paper 98-1, 140 p., available online at 

http://www.nrcs.usda.gov/technical/land/pubs/manntr.pdf 

Kemp, W.M., Sampou, P., Caffrey, J., and Mayer, M., 1990, Ammonium recycling 

versus denitrification in Chesapeake Bay sediments: Limnology & Oceanography, 

v. 351, p. 545–563. 

Khan, S.A., Mulvaney, R., and Hoeft, R.G., 2001, A simple soil test for detecting sites 

that are nonresponsive to nitrogen fertilization: Soil Science Society of America 

Journal, v. 65, p. 1751–1760. 
248

Khanna, M., 2001, Sequential adoption of site-specific technologies and its implications 

for nitrogen productivity: A double selectivity model: American Journal of 

Agricultural Economics, v. 83, p. 35-51. 

Khanna, M., Isik, M., and Winter-Nelson, A., 2000, Investment in site-specific crop 

management under uncertainty: Implications for nitrate pollution control and 

environmental policy: Agricultural Economics, v. 24, no. 1, p. 9-21. 

Khanna, M., Yang, W., Farnsworth, R., and Onal, H., 2003, Cost effective targeting of 

CREP to improve water quality with endogenous sediment deposition coefficients: 

American Journal of Agricultural Economics, v. 85, p. 538–553. 

Khoshmanesh, A., Hart, B.T., Duncan, A., and Beckett, R., 1999, Biotic uptake and 

release of phosphorus by a wetland sediment: Environmental Technology, v. 29, 

p. 85–91. 

King, D., 2005, Crunch time for water quality trading: Choices, v. 20, p. 71–75. 

Kirsch, K., Kirsch, A., and Arnold, J.G., 2002, Predicting sediment and phosphorus loads 

in the Rock River Basin using SWAT: Transactions of the American Society of 

Agricultural and Biological Engineers, v. 45, no. 6, p. 1757–1769. 

Kitchen, N.R., Sudduth, K.A., and Drummond, S.T., 1999, Soil electrical conductivity as 

a crop productivity measure for claypan soils: Journal of Production Agriculture, 

v. 12, p. 607–617. 

Kladivko, E.J., Frankenberger, J.R., Jaynes, D.B., Meek, D.W., Jenkinson, B.J., and 

Fausey, N.R., 2004, Nitrate leaching to subsurface drains as affected by drain 

spacing and changes in crop production system: Journal of Environmental Quality, 

v. 33, p. 1803–1813. 

Kling, C., Secchi, S., Jha, M., Feng, H., Gassman, P., and Kurkalova, L., 2006, Upper 

Mississippi River basin modelling system, Part 3—Conservation practice scenario 

results, in Singh, V., and Xu. Y., eds., Water Resources Publication, Coastal 

Hydrology and Processes, p. 127-134. 

Koch, B., Khosla, R., Frasier, W.M., Westfall, D.G., and Inman, D., 2004, Economic 

feasibility of variable-rate nitrogen application utilizing site-specific management 

zones: Agronomy Journal, v. 96, p. 1572–1580. 

Kovacic, D.A., David, M.B., Gentry, L.E., Starks, K.M., and Cooke, R.A., 2000, 

Effectiveness of constructed wetlands in reducing nitrogen and phosphorus export 

from agricultural tile drainage: Journal of Environmental Quality, v. 29, p. 1262–

1274. 

Krom, M.D., and Berner, R.A., 1980, Adsorption of phosphate in anoxic marine 

sediments: Limnology & Oceanography, v. 25, p. 797–806.

Laboski, C.A.M., Sawyer, J.E., Walters, D.T., Bundy, L.G., Hoeft, R.G., Randall, G.W., 

and Andraski, T.W., 2006, Evaluation of the Illinois soil nitrogen test in the north 

central region, in Proceedings of the North Central Extension-Industry Soil Fertility 

Conference, Des Moines, IA, v. 22, p. 86–93. 
249

Lambert, D.M., Lowenberg-DeBoer, J., and Malzer, G.L., 2006, Economic analysis of 

spatial-temporal patterns in corn and soybean response to nitrogen and phosphorus: 

Agronomy Journal, v. 98, p. 43–54. 

Landeck-Miller, R.E., and St. John, J.P., 2006, Modeling primary production in the 

Lower Hudson River Estuary, in Levington, J.S., and Waldman, J.R., eds., The 

Hudson River Estuary, New York, Cambridge University Press, p. 140-153. 

Lane, R.R., Day, J.W., Marx, B., Reyes, E., and Kemp, G.P., 2002, Seasonal and spatial 

water quality changes in the outflow plume of the Atchafalaya River, Louisiana, 

USA: Estuaries, v. 25, no. 1, p. 30–42. 

Lanyon, L.E., 2005, Phosphorus, animal nutrition and feeding: Overview, in Sims, J.T., 

and Sharpley, A.N., eds., Phosphorus: Agriculture and the Environment, Madison, 

WI, American Society of Agronomy Monograph Series No. 46, p. 561-586. 

Larson, R.S., 2001, Water quality trends of the Illinois waterway system upstream of 

Peoria including the Chicago metropolitan area: Champaign, IL, Illinois State Water 

Survey, Contract Report 2001–03. 

Leavitt, P.R., and Hodgson, D.A., 2001, Sedimentary pigments, in Smol, J.P., Birks, 

H.J.B, and Last, W.M., eds., Tracking environmental change using lake sediments, 

v. 3, Terrestrial, algal, and siliceous indicators: Kluwer Academic Publishers, 

p. 295–325. 

Lee, K.H., Isenhart, T.M., and Schultz, R.C., 2003, Sediment and nutrient removal in an 

established multi-species riparian buffer: Journal of Soil and Water Conservation, 

v. 58, no. 1, p. 1–7. 

Lee, K.H., Isenhart, T.M., Schultz, R.C., and Mickelson, S.K., 2000, Multispecies 

riparian buffers trap sediment and nutrients during rainfall simulations: Journal of 

Environmental Quality, v. 29, no. 4, p. 1200–1205. 

Lemunyon, J.L., and Gilbert, R.G., 1993, The concept and need for a phosphorus 

assessment tool: Journal of Production Agriculture, v. 6, p. 483–496. 

Lewandrowski, J., Peters, M., Jones, C., House, R., Sperow, M., Eve, M., and Paustian, 

K., 2004, Economics of sequestering carbon in the U.S. agricultural sector: U.S. 

Department of Agriculture, Economic Research Service, Technical Bulletin 

No. 1909, 69 p. 

Liu, H.B., Dagg, J.M. Campbell, L., and Urban-Rich, J., 2004, Picophytoplankton and 

bacterioplankton in the Mississippi River plume and its adjacent waters: Estuaries, 

v. 27, no. 1, p. 147–156. 

Lohrenz, S.E., Dagg, M.L., and Whitledge, T.E., 1990, Enhanced primary production at 

the plume/oceanic interface of the Mississippi River: Continental Shelf Research, 

v. 10, no. 7, p. 639–664.

Lohrenz, S.E., Fahnenstiel, G.L., and Redalje, D.G., 1994, Spatial and temporal 

variations of photosynthetic parameters in relation to environmental conditions in 

northern Gulf of Mexico coastal waters: Estuaries, v. 17, p. 779–795.
250

Lohrenz, S.E., Fahnenstiel, G.L., Redalje, D.G., and Lang, G.A., 1992, Regulation and 

distribution of primary production in the northern Gulf of Mexico, in Program, 

N.C.O., ed., Nutrient Enhanced Coastal Ocean Productivity, NECOP Workshop 

Proceedings, October 1991: College Station, TX, Texas Sea Grant Publications, 

p. 95–104.

Lohrenz, S.E., Fahnenstiel, G.L., Redalje, D.G., Lang, G.A., Chen, X.G., and Dagg, M.J., 

1997, Variations in primary production of northern Gulf of Mexico continental shelf 

waters linked to nutrient inputs from the Mississippi River: Marine Ecology-Progress 

Series, v. 155, p. 45–54.

Lohrenz, S.E., Fahnenstiel, G.L., Redalje, D.G., Lang, G.A., Dagg, M.J., Whitledge, 

T.E., and Dortch, Q., 1999a, Nutrients, irradiance, and mixing as factors regulating 

primary production in coastal waters impacted by the Mississippi River plume: 

Continental Shelf Research, v. 19, p. 1113–1141.

Lohrenz, S.E., Wiesenburg, D.A., Arnone, R.A., and Chen, X.G., 1999b, What controls 

primary production in the Gulf of Mexico?, in Kumpf, H., Steidinger, K., and 

Sherman, K., eds., The Gulf of Mexico large marine ecosystem: Malden, MA, 

Assessment, Sustainability and Management, Blackwell Science, Inc., p. 151–170.

Lory, J.A., and Scharf, P.C., 2003, Yield goal versus delta yield for predictiong fertilizer 

nitrogen need in corn: Agronomy Journal, v. 95, p. 994-999. 

Loureiro, M.L., McCluskey, J.J., and Mittelhammer, R.C., 2001, Assessing consumers 

preferences for organic, eco-labeled and regular apples: Journal of Agricultural and 

Resource Economics, v. 26, no. 2, p. 404–416. 

Lubowski, R.N., Bucholtz, S., Claassen, R., Roberts, M.J., Cooper, J.C., Gueorguieva, 

A., and Johansson, R., 2006, Environmental effects of agricultural land-use change: 

The role of Economics and policy: U.S. Department of Agriculture, Economic 

Research Service Report Number ERR-25, 82 p., available online at: 

http://www.ers.usda.gov/Publications/ERR25/. 

LUMCON, 2007, Dead zone size near top end: LUMCON News, available on line at: 

http://www.lumcon.edu/Information/news/default.asp?XMLFilename=20070731164

8.xml, last accessed August 7, 2007. 

MacKinnon, J.A., and Gregg, M.C., 2005, Spring mixing: Turbulence and internal waves 

during restratification on the New England shelf: Journal of Physical Oceanography, 

v. 35, p. 2425–2443. 

Mallarino, A.P., and Schepers, J.S., 2005, Role of precision agriculture in phosphorus 

management practices, in Sims, J.T., and Sharpley, A.N., eds., Phosphorus: Madison, 

WI, Agriculture and the Environment, American Society of Agronomy Monograph 

Series No. 46, Crop Science Society of America and Soil Science Society of 

America, p. 881–908. 

Maloney, M., and Brady, G., 1988, Capital turnover and marketable emission rights: 

Journal of Law and Economics, v. 31, p. 203–226. 
251

Mamo, M., Malzer, G.L., Mulla, D.J., Huggins, D.R., and Strock, J., 2003, Spatial and 

temporal variation in economically optimum nitrogen rate for corn: Agronomy 

Journal, v. 95, p. 958-964. 

Mann, L., and Tolbert, V., 2000, Soil sustainability in renewable biomass plantings: 

Ambio, v. 29, no. 8, p. 492–498. 

Marquez, C.O., Cambardella, C.A., Isenhart, T.M., and Schultz, R.C., 1999, Assessing 

soil quality in a riparian buffer strip system by testing organic matter fractions: 

Agroforestry Systems, v. 44, p. 133–140. 

MART, 2006a, Management Action Review Team Report: U.S. Environmental 

Protection Agency, Mississippi River/Gulf of Mexico Watershed Nutrient Task 

Force, Management Action Reassessment Team, 31 p., available online at 

http://www.epa.gov/msbasin/taskforce/MART.pdf. 

MART, 2006b, Reassessment of point source nutrient mass loadings to the Mississippi 

River basin: U.S. Environmental Protection Agency, Mississippi River/Gulf of 

Mexico Watershed Nutrient Task Force, Management Action Reassessment Team, 

31 p., available online at 

http://www.epa.gov/msbasin/taskforce/Point_Source_Mass_Loading.pdf. 

Maryland Department of the Environment, 2005, Chesapeake Bay Restoration Program: 

Baltimore, MD. 

Mayer, P.M., Reynolds, S.K., McCutchen, M.D., and Canfield, T.J., 2006, Riparian 

buffer width, vegetative cover, and nitrogen removal effectiveness—A review of 

current science and regulations: Cincinnati, OH, U.S. Environmental Protection 

Agency, EPA/600/R-05/118. 

Mazurek, J., 2002, Government-sponsored voluntary programs for firms—An initial 

survey, in Dietz, T., and Stern, P.C., eds., New Tools for Environmental Protection: 

Education, Information, and Voluntary Measures, The National Academies Press, 

National Academy of Sciences, p. 219-234. 

McDowell, L.L., and McGregor, K.C., 1984, Plant nutrient losses in runoff from 

conservation tillage corn: Soil Tillage Research, v. 4, p. 79–91. 

McDowell, R.W., and Sharpley, A.N., 2001, A comparison of fluvial sediment 

phosphorus (P) chemistry in relation to location and potential to influence stream P 

concentrations: Aquatic Geochemistry, v. 7, p. 255–265. 

McDowell, R.W., and Sharpley, A.N., 2003, Uptake and release of phosphorus from 

overland flow in a stream environment: Journal of Environmental Quality, v. 32, 

p. 937–948. 

McDowell, R.W., Sharpley, A.N., and Kleinman, P.J.A., 2002, Integrating phosphorus 

and nitrogen decision management at watershed scales: Journal of American Water 

Resources Association, v. 38, no. 2, p. 479–491. 

McGraw, T., and Hemb, R., 1995, Fertility variability in the Minnesota River Valley 

watershed in 1993 as determined from grid soil testing results on 52,000 acres on 

commercial fields, in Robert, P.C., Rust, R.H., and Larson, W.E., eds., Site-specific 
252

management for agricultural systems: Minneapolis, MN, American Society of 

Agronomy, Crop Science Society of America and Soil Science Society of America, 

2nd International Conference. 

McIsaac, G.F., 2006, Net anthropogenic nitrogen inputs (NANI) to the Mississippi River 

basin: Presented at Science Symposium—Sources, transport, and fate of nutrients in 

the Mississippi and Atchafalaya River basins, November 7–9, 2006, Minneapolis, 

MN. 

McIsaac, G.F., David, M.B., Gertner, G.Z., and Goolsby, D.A., 2001, Nitrate flux in the 

Mississippi River: Nature, v. 414, p. 166–167. 

McIsaac, G.F., David, M.B., Gertner, G.Z., and Goolsby, D.A., 2002, Relating net 

nitrogen input in the Mississippi River basin to nitrate flux in the lower Mississippi 

River—A comparison of approaches: Journal of Environmental Quality, v. 31, 

p. 1610–1622. 

McIsaac, G.F., and Hu, X., 2004, Net N input and riverine N export from Illinois 

agricultural watersheds with and without extensive tile drainage: Biogeochemistry, 

v. 70, p. 251–271. 

McKee, B.A., Aller, R.C., Allison, M.A., Bianchi, T.S., and Kineke, G.C., 2004. 

Transport and transformation of dissolved and particulate materials on continental 

margins influenced by major rivers—Benthic boundary layer and seabed processes: 

Continental Shelf Research, v. 24, p. 899–926. 

McLaughlin, S.B., and Kszos, L.A., 2005, Biomass and bioenergy development of 

switchgrass (Panicum virgatum) as a bioenergy feedstock in the United States: 

Biomass & Bioenergy, v. 28, p. 515–535.

McLaughlin, S., and Walsh, M., 1998, Evaluating environmental consequences of 

producing herbaceous crops for bioenergy: Biomass & Bioenergy, v. 14, no. 4, 

p. 317–324. 

McMahon, G., Alexander, R.B., and Qian, S., 2003, Support of total maximum daily load 

programs using spatially referenced regression models: Journal of Water Resources 

Planning and Management, v. 129, p. 315–329. 

Mee, L.D., 2006, Reviving dead zones: Scientific American, v. 295, p. 78–85.

Meisinger, J.J., and Delgado, J.A., 2002, Priciples for managing nitrogen leaching: 

Journal of Soil and Water Conservation, v. 57, no. 6, p. 485-499. 

Melillo, J.M., and Cowling, E.B., 2002, Reactive nitrogen and public policies for 

environmental protection: AMBIO, A Journal of the Human Environment, v. 31, 

no. 2, p. 141–149. 

Mellor, G.L., and Yamada, T., 1982, Development of a turbulence closure model for 

geophysical fluid problems: Reviews of Geophysics and Space Physics, v. 20, 

p. 851–875. 

Meretsky, V.J., Wegner, D.L., and Stevens, L.E., 2000, Balancing endangered species 

and ecosystems—A case study of adaptive management in Grand Canyon: 

Environmental Management, v. 25, no. 6, p. 579–586. 
253

Metropolitan Council, 2004, Regional progress in water quality—Analysis of water 

quality data from 1976 to 2002 for the major rivers in the Twin Cities: St. Paul, MN, 

Metropolitan Council, Publication 32-04-045. 

Mitsch, W.J., Day, J.W., Jr., Gilliam, W., Groffman, P.M., Hey, D.L., Randall, G.W., and 

Wang, N., 2001, Reducing nitrogen loading to the Gulf of Mexico from the 

Mississippi River basin—Strategies to counter a persistent ecological problem: 

Bioscience, v. 51, p. 373–388.

Mitsch, W.J., Day, J.W., Jr., Gilliam, J.W., Groffman, P.M., Hey, D.L., Randall, G.W., 

and Wang, N., 1999, Reducing nutrient loads, especially nitrate-nitrogen, to surface 

water, ground water, and the Gulf of Mexico: Topic 5 report for the integrated 

assessment on hypoxia in the Gulf of Mexico: Silver Spring, MD, National Oceanic 

and Atmospheric Administration Coastal Ocean Program Decision Analysis Series 

No. 19, 111 p. Available on line at:

http://oceanservice.noaa.gov/products/hypox_t5final.pdf. 

Mitsch, W.J., Day, J.W., Jr., Zhang, L., and Lane, R.R., 2005a, Nitrate-nitrogen retention 

in wetlands in the Mississippi River Basin: Ecological Engineering, v. 24, p. 267–

278. 

Mitsch, W.J., Zhang, L., Anderson, C.J., Altor, A.E., and Hernandez, M.E., 2005b, 

Creating riverine wetlands—Ecological succession, nutrient retention, and pulsing 

effects: Ecological Engineering, v. 25, p. 510–527. 

Montagna, P., Hodges, B., Maidment, D., and Minsker, B., 2006, Long-term studies of 

hypoxia in Corpus Christi Bay: The cybercollaboratory testbed: Paper presented at 

Hypoxia Effects on Living Resources in the Gulf of Mexico, September 25–26, 

2006: New Orleans, Louisiana, Tulane University, sponsored by National Oceanic 

and Atmospheric Administration Center for Sponsored Coastal Ocean Research. 

Moomaw, W.R., 2002, Energy, industry and nitrogen—Strategies for decreasing reactive 

nitrogen emissions: AMBIO, A Journal of the Human Environment, v. 31, no. 2, 

p. 184–189. 

Morey, S.L., Martin, P.J., O’Brien, J.J., Wallcraft, A.A., Zavala-Hidalgo, J., 2003a, 

Export pathways for river discharged fresh water in the northern Gulf of Mexico: 

Journal of Geophysical Research, v. 108, C10, 3303, doi:1029/2002JC001674.

Morey, S.L., Schroeder, W.W., O’Brien, J.J., and Zavala-Hidalgo, J., 2003b, The annual 

cycle of riverine influence in the eastern Gulf of Mexico basin: Geophysical 

Research Letters, v. 30, 1867, doi:10.1029/2003GL017348.

Morgenstern, R., and Pizer, W., eds., 2007, Reality check—The nature and performance 

of voluntary environmental programs in the United States, Europe, and Japan: 

Washington, D.C., Resources for the Future Press, 200 p. 

Morse, J.W., and Eldridge, P.M., in press, A non-steady state diagenetic model for 

changes in sediment biogeochemistry in response to seasonally hypoxic/anoxic 

conditions in the “dead zone” of the Louisiana shelf: Marine Chemistry, v. __, p. __–

__, Accepted for publication, draft available online at: 
254

http://yosemite.epa.gov/sab/sabhap.nsf/2a890dc663b46bc685256d63006ac3aa/6ca36

cf055bc5b2d8525727c004bc993/$FILE/Morse%20and%20Eldridge%202007.pdf.

Morse, J.W., and Rowe, G.T., 1999, Benthic biogeochemistry beneath the Mississippi 

River plume: Estuaries, v. 22, p. 206–214.

Mortimer, C.H.J., 1941, The exchange of dissolved substances between mud and water in 

lakes—I and II: Journal of Ecology, v. 29, p. 280–329. 

Mosier, A., Kroeze, C., Nevison, C., Oenema, O., Seitzinger, S., and van Cleemput, O., 

1998, Closing the global N2O budget: Nitrous oxide emissions through the 

agricultural nitrogen cycle: Nutrient Cycling in Agroecosystems, v. 52, no. 2-3, p. 

225-248, available on-line at: 

http://www.springerlink.com/content/lu04703848m261w5/. 

MR/GMWNTF, 2001, Action plan for reducing, mitigating, and controlling hypoxia in 

the northern Gulf of Mexico: Washington, D.C., Mississippi river/Gulf of Mexico 

Watershed Nutrient Task Force, 36 p. Available on line at: 

http://www.epa.gov/msbasin/taskforce/pdf/actionplan.pdf.

Mueller, D.H., Wendt, R.C., and Daniel, T.C., 1984, Phosphorus losses as affected by 

tillage and manure application: Soil Science Society of America Journal, v. 48, 

p. 901–905. 

Mulholland, P.J., Steinman, A.D., Marzolf, E.R., Hart, D.R., and DeAngelis, D.L., 1994, 

Effect of periphyton biomass on hydraulic characteristics and nutrient cycling in 

streams: Oecologia, v. 98, p. 40–47. 

Mulvaney, R.L., Khan, S.A., Hoeft, R.G., and Brown, H.M., 2001, A soil organic 

nitrogen fraction that reduces the need for nitrogen fertilization: Soil Science Society 

of America Journal, v. 65, p. 1164–1172. 

Murdock, L.W., Howe, P.L., and Schwab, G.J., 2002, Variable rate nitrogen fertilizer for 

corn grown in Kentucky, in Proceedings of the North-Central Extension-Industry 

Soil Fertility Conference: Brookings, SD, Potash and Phosphate Institute, v. 18, 

p. 81–85. 

NASA-SeaWiFS, 2007, Sea-viewing Wide Field-of-view Sensor (SeaWiFS) Project: 

National Aeronautics and Space Administration, available on line at: 

http://oceancolor.gsfc.nasa.gov/SeaWiFS/. 

National Academy of Sciences, 2007, Mississippi River water quality and the Clean 

Water Act: Progress, challenges, and opportunities: National Research Council, 

National Academy of Sciences, Committee on the Mississippi River and the Clean 

Water Act, 215 p., available on line at: 

http://www.nap.edu/catalog.php?record_id=12051.

National Research Council, 2000, Clean coastal waters—Understanding and reducing the 

problems from nutrient pollution: Washington, D.C., National Academy of Sciences 

Press, 405 p. 

National Research Council, 2004, Adaptive management for water resource planning: 

Washington, D.C., National Academy of Science Press, 123 p. 
255

Neal, C., 2001, The potential for phosphorus pollution mediation by calcite precipitation 

in UK freshwaters: Hydrology and Earth System Sciences, v. 5, p. 119–131. 

Neitsch, S.L., Arnold, J.G., Kiniry, J.R., Srinivasan, R., and Williams, J.R., 2004, Soil 

and water assessment tool input/output file documentation, version 2005: Temple, 

TX, U.S. Department of Agriculture, Agricultural Research Service, Grassland, Soil 

and Water Research Laboratory, available online at: 

ftp://ftp.brc.tamus.edu.pub/outgoing/sammons/swat2005 (accessed 11/28/06). 

Nelson, R., Tietenberg, T., and Donihue, M.R., 1993, Differential environmental 

regulation: effects on electric utility capital turnover and emissions: Review of 

Economics and Statistics, v. 75, p. 368–373. 

Nielsen, K., Nielsen, L.P., and Rasmussen, P., 1995, Estuarine nitrogen retention 

independently estimated by the denitrification rate and mass balance methods—A 

study of Norsminde Fjord, Denmark: Marine Ecology Progress Series, v. 119, 

p. 273–283.

Nixon, S.W., 1992, Quantifying the relationship between nitrogen input and the 

productivity of marine ecosystems: Tokyo, Japan, Advanced Marine Technology 

Conference, no. 5.

Nixon, S.W., Ammerman, J., Atkinson, L., Berounsky, V., Billen, G., Boicourt, W., 

Boynton, W., Church, T., DiToro, D., Elmgren, R., Garber, J., Giblin, A., Jahnke, R., 

Owens, N., Pilson, M.E.Q., and Seitzinger, S., 1996, The fate of nitrogen and 

phosphorus at the land-sea margin of the North Atlantic Ocean: Biogeochemistry, 

v. 35, p. 141–180.

Nixon, S.W., Kelly, J.R., Furnas, B.N., Oviatt, C.A., and Hale, S.S., 1980, Phosphorus 

regeneration and the metabolism of coastal marine bottom communities, in Tenore, 

K.R., and Coull, B.C., eds., Marine Benthic Dynamics: University of South Carolina 

Press, p. 219–242.

NOAA, 2007, Nutrient Enhanced Coastal Ocean Productivity (NECOP) Program: 

National Oceanic and Atmospheric Administration, Washington, DC, available 

online at: http://www.aoml.noaa.gov/ocd/necop/, accessed April, 2007. 

Northcott, W.J., Cooke, R.A., Walker, S.E., Mitchell, J.K., and Hirschi, M.C., 2001, 

Application of DRAINMOD-N to fields with irregular drainage systems: 

Transactions of the American Society of Agricultural Engineers, v. 44, no. 2, p. 241-

249. 

Novak, J.M., Stone, K.C., Watts, D.W., and Johnson, M.H., 2003, Dissolved phosphorus 

transport during storm and base flow conditions from an agriculturally intensive 

Southeastern Coastal Plain watershed: Transactions of the American Society of 

Agricultural Engineers, v. 46, no. 5, p. 1355–1363. 

Nowlin, W.D., Jochems, A.E., DiMarco, S.F., Reid, R.O., and Howard, M.K., 2005, 

Low-frequency circulation over the Texas-Louisiana continental shelf, in Sturges, 

W., and Lugo-Fernandez, A., eds., Circulation in the Gulf of Mexico—Observations 

and models: American Geophysical Union, Geophysical Monograph, v. 161, p. 219–

240. 
256

Oberle, S.L., and Keeney, D.R., 1990, Soil type, precipitation, and fertilizer N effects on 

corn yields: Journal of Production Agriculture, v. 3, p. 522–527. 

O’Connor, T., and Whitall, D., 2007, Linking hypoxia to shrimp catch in the northern 

Gulf of Mexico: Marine Pollution Bulletin, v. 54, no. 4, p. 460-463. 

O’Donnell, T.K., and Galant, D.L., in press, Evaluating success criteria and project 

monitoring in river enhancement within an adaptive management framework: 

Environmental Management, (in press). 

Oey, L.-Y., Ezer, T., Forristall, G., Cooper, C., DiMarco, S., and Fan, S., 2005a, An 

exercise in forecasting loop current and eddy frontal positions in the Gulf of Mexico: 

Geophysical Research Letters, v. 32, L12611, doi:10.1029/2005GL023253. 

Oey, L.-Y., Ezer, T., and Lee, H.-C., 2005b, Loop current rings and related circulation in 

the Gulf of Mexico—A review of numerical models and future challenges, in

Sturges, W., and Fernandez, A., eds., Circulation in the Gulf of Mexico—

Observations and models: American Geophysical Union, p. 31–56.

Ogden, J.C., Davis, S.M., Jacobs, K.J., Barnes, T., and Fling, H.E., 2005, The use of 

conceptual ecological models to guide ecosystem restoration in south Florida: 

Wetlands, v. 25, no. 4, p. 795–809.

Oguz, T., and Gilbert, D., 2007, Abrupt transitions of the top-down controlled Black Sea 

pelagic ecosystem during 1960–2000—Evidence for regime-shifts under strong 

fishery exploitation and nutrient enrichment modulated by climate-induced 

variations: Deep-Sea Research, v. 54, p. 220–242. 

Ohlmann, J.C., and Niiler, P.P., 2005, Circulation over the continental shelf in the 

northern Gulf of Mexico: Progress in Oceanography, v. 64, p. 45–81. 

Olness, A.E., Smith, S.J., Rhoades, E.D., and Menzel, R.G., 1975, Nutrient and sediment 

discharge from agricultural watersheds in Oklahoma: Journal of Environmental 

Quality, v. 4, p. 331–336. 

Olson, R.J., Hensler, R.F., Attoe, O.J., Witzel, S.A., and Peterson, L.A., 1970, Fertilizer 

nitrogen and crop rotation in relation to movement of nitrate nitrogen through soil 

profiles: Soil Science Society of America Proceedings, v. 34, p. 448–452. 

Omay, A.B., Rice, C.W., Maddux, L.D., and Gordon, W.B., 1997, Changes in soil 

microbial and chemical properties under long-term crop rotation and fertilization: 

Soil Science Society of America Journal, v. 61, p. 1672–1678. 

Opdyke, M.R., David, M.B., and Rhoads, B.L., 2006, The influence of geomorphological 

variability in channel characteristics on sediment denitrification in agricultural 

streams: Journal of Environmental Quality, v. 35, p. 2103–2112. 

Osborne, S.L., Schepers, J.S., Francis, D.D., and Schlemmer, M.R., 2002, Detection of 

phosphorus and nitrogen deficiencies in corn using spectral radiance measurements: 

Agronomy Journal, v. 94, p. 1215–1221. 

Osterman, L.E., Poore, R.Z., Swarzenski, P.W., and Turner, R.E., 2005, Reconstructing a 

180-yr record of natural and anthropogenic induced hypoxia from the sediments of 

the Louisiana continental shelf: Geology, v. 33, no. 4, p. 329–332.
257

Owens, L.B., Edwards, W.M., and Van Keuren, R.W., 1997, Runoff and sediment losses 

resulting from winter feeding on pastures: Journal of Soil and Water Conservation, 

v. 52, p. 194–197. 

Owens, L.B., Malone, R.W., Shiptalo, M.J., Edwards, W.M., and Bonta, J.V., 2000, 

Lysimeter study of nitrate leaching from a corn-soybean rotation: Journal of 

Environmental Quality, v. 29, p. 467-474. 

Paerl, H.W., and Fulton, R.S., III, 2006, Ecology of harmful cyanobacteria, in Graneli, 

E., and Turner, J., eds., Ecology of harmful marine algae: Berlin, Springer-Verlag, 

p. 69–78.

Paerl, H.W., Valdes, L.M., Piehler, M.F., and Stow, C.A., 2006b, Assessing the effects of 

nutrient management in an estuary experiencing climatic change—The Neuse River 

Estuary, NC, USA: Environmental Management, v. 37, p. 422–436.

Pakulski, J.D., Benner, R., Whitledge, T., Amon, R., Eadie, B., Cifuentes, L., 

Ammerman, J., and Stockwell, D., 2000, Microbial metabolism and nutrient cycling 

in the Mississippi and Atchafalaya River plumes: Estuarine and Coastal Shelf 

Science, v. 50, no. 2, p. 173–184. 

Paludan, C., and Blicher-Mathiesen, G., 1996, Losses of inorganic carbon and nitrous 

oxide from a temperate freshwater wetland in relation to nitrate loading: 

Biogeochemistry, v. 35, no. 2, p. 305-326. 

Parkin, T.B., and Kaspar, T.C., 2006, Nitrous oxide emissions from corn-soybean 

systems in the Midwest: Journal of Environmental Quality, v. 35, p. 1496-1506.

Patrick, W.H., Jr., Gotoh, S., and Williams, B.G., 1973, Strengite dissolution in flooded 

soils and sediments: Science, v. 179, p. 564–565. 

Patterson, B.D., 1987, The principle of nested subsets and its implications for biological 

conservation: Conservation Biology, v. 1, p. 323–334. 

Pavelis, G.A., ed., 1987, Farm drainage in the United States—History, status, and 

prospects: U.S. Department of Agriculture, Economic Research Service, Misc. Pub. 

1455, 170 p. 

Perez, A.N., Oehlers, L., and Walter, R.B., 2006, Detection of hypoxia-related proteins in 

Medaka (Oryzias latipes) by difference gel electrophoresis and identification by 

sequencing of peptides using MALDI-TOF mass spectrometry: Paper presented at 

Hypoxia Effects on Living Resources in the Gulf of Mexico, September 25–26, 

2006, Tulane University, New Orleans, Louisiana, sponsored by National Oceanic 

and Atmospheric Administration Center for Sponsored Coastal Ocean Research. 

Perlack, R.D., and Turhollow, A.F., 2003, Feedstock cost analysis of corn stover residues 

for further processing: Energy, v. 28, no. 14, p. 1395-1403. 

Perlack, R.D., Wright, L.L., Turhollow, A.F., Graham, R.L., Stokes, B.J., Erbach, D.C., 

2005, Biomass as feedstock for a bioenergy and bioproducts industry: The technical 

feasibility of a billion-ton annual supply: Oak ridge National Laboratory, 78 p., 

available on line at: http://feedstockreview.ornl.gov/pdf/billion_ton_vision.pdf. 
258

Peters, R.H., 1986, The role of prediction in limnology: Limnology and Oceanography, 

v. 31, p. 1143–1159.

Phipps, R.G., 1997, Nitrate removal capacity of constructed wetlands: Ames, IA, Iowa 

State University, Ph.D. dissertation, 68 p. 

Phipps, R.G., and Crumpton, W.G., 1994, Factors affecting nitrogen loss in experimental 

wetlands with different hydrologic loads: Ecological Engineering, v. 3, p. 399–408. 

Pierson, S.T., Cabrere, M.L., Evanylo, G.K., Kuykendall, H.A., Hoveland, C.S., 

McCann, M.A., and West, L.T., 2001, Phosphorus and ammonium concentrations in 

surface runoff from grasslands fertilized with broiler litter: Journal of Environmental 

Quality, v. 30, p. 1784–1789. 

Platon, E., and Sen Gupta, B.K., 2001, Benthic foraminiferal communities in oxygen 

depleted environments of the Louisiana Continental Shelf, in Rabalais, N.N., and 

Turner, R.E., eds., Coastal hypoxia—Consequences for living resources and 

ecosystems: American Geophysical Union, Coastal and Estuarine Study Series, 

v. 58, p. 147–163.

Platon, E., Sen Gupta, B.K., Rabalais, N.N., and Turner, R.E., 2005, Effect of seasonal 

hypoxia on the benthic forminiferal community of the Louisiana inner continental 

shelf—The 20th century record: Marine Micropaleontology, v. 54, p. 263–283. 

Popova, Y.A., Keyworth, V.G., Haggard, B.E., Storm, D.E., Lynch, R.A., and Payton, 

M.E., 2006, Stream nutrient limitation and sediment interactions in the EuchaSpavinaw basin, USA: Journal of Soil and Water Conservation, v. 61, p. 105–115. 

Pote, D.H., Daniel, T.C., Nichols, D.J., Sharpley, A.N., Moore, P.A., Jr., Miller, D.M., 

and Edwards, D.R., 1999, Relationship between phosphorus levels in three Ultisols 

and phosphorus concentrations in runoff: Journal of Environmental Quality, v. 28, 

p. 170–175. 

Pote, D.H., Daniel, T.C., Sharpley, A.N., Moore, P.A., Jr., Edwards, D.R., and Nichols, 

D.J., 1996, Relating extractable soil phosphorus to phosphorus losses in runoff: Soil 

Science Society of America Journal, v. 60, p. 855–859. 

Power, J.F., Wiese, R., and Flowerday, D., 2000, Managing nitrogen for water quality—

Lessons from management systems evaluation area: Journal of Environmental 

Quality, v. 29, p. 355–366. 

PPI/PPIC/FAR, 2005, Soil test levels in North America -- Summary update: Norcross, 

GA, Potash & Phosphate Institute, PPI/PPIC/FAR Technical Bulletin 2005-1. 

PPI/PPIC/FAR, 2002, Plant nutrient use in North American agriculture—Producing food 

and fiber, preserving the environment and integrating organic and inorganic sources: 

Norcross, GA, Potash & Phosphate Institute, PPI/PPIC/FAR Technical Bulletin 

2002-1. 

Preston, S.D., and Brakebill, J.W., 1999, Application of spatially referenced regression 

modeling for the evaluation of total nitrogen loading in the Chesapeake Bay 

watershed: U.S. Geological Survey Water-Resources Investigations Report 99-4054, 
259

12 p., available online at http://md.water.usgs.gov/publications/wrir-99-

4054/html/index.htm

Prospero, J.M., Barrett, K., Church, T., Dentener, F., Duce, R.A., Galloway, J.N., Levy, 

H., Moody, J., and Quinn, P., 1996, Atmospheric deposition of nutrients to the North 

Atlantic Basin: Biogeochemistry, v. 35, p. 27–73. 

Rabalais, N.N., 2006, Benthic communities and the effects of hypoxia in Louisiana 

coastal waters: Paper presented at Hypoxia Effects on Living Resources in the Gulf 

of Mexico, September 25–26, 2006, Tulane University, New Orleans, Louisiana, 

sponsored by National Oceanic and Atmospheric Administration Center for 

Sponsored Coastal Ocean Research. 

Rabalais, N.N., Atilla, N., Normandeau, C., and Turner, R.E., 2004, Ecosystem history of 

Mississippi River-influenced continental shelf revealed through preserved 

phytoplankton pigments: Marine Pollution Bulletin, v. 49, p. 537–547. 

Rabalais, N.N., Lohrenz, S.E., Redalje, D.G., Dortch, Q., Justić, D., Turner, R.E., 

Qureshi, N.A., Dagg, M.J., Eadie, B.J., and Fahnensteil, G.L., 1999b, Nutrientenhanced coastal productivity and ecosystem responses, in Wiseman, W.J., Jr., 

Rabalais, N.N., Dagg, M.J., and Whitledge, T.E., eds., chap. 4, Nutrient enhanced 

coastal ocean productivity in the northern Gulf of Mexico: Silver Spring, MD, 

National Oceanic and Atmospheric Administration Coastal Ocean Program, Decision 

Analysis Series No. 14, U.S. Department of Commerce, National Ocean Service, 

Center for Sponsored Coastal Research, p. 51–78. 

Rabalais, N.N., and Turner, R.E., 2001, Hypoxia in the northern Gulf of Mexico—

Description, causes, and change, in Rabalais, N.N., and Turner, R.E., eds., Coastal 

hypoxia— Consequences for living resources and ecosystems: Washington, D.C., 

American Geophysical Union, Coastal and Estuarine Studies, v. 58, 454 p. 

Rabalais, N.N., and Turner, R.E., 2006, Oxygen depletion in the Gulf of Mexico adjacent 

to the Mississippi River, in Neretin, L.N., ed., Past and present marine water column 

anoxia: NATO Science Series, IV-Earth and Environmental Sciences, Kluwer, 

p. 225–245. 

Rabalais, N.N., Turner, R.E., Dortch, Q., Justić, D., Bierman, V.J., and Wiseman, W.J., 

2002, Nutrient-enhanced productivity in the northern Gulf of Mexico—Past, present 

and future: Hydrobiologia, v. 475, p. 39–63. 

Rabalais, N.N., Turner, R.E., Justić, D., Dortch, Q., and Wiseman, W.J., 1999a, 

Characterization of hypoxia: Topic 1 report for the integrated assessment of hypoxia 

in the Gulf of Mexico: Silver Spring, MD, National Oceanic and Atmospheric 

Administration Coastal Ocean Program Decision Analysis Series No. 15, 203 p. 

Available on line at: http://oceanservice.noaa.gov/products/hypox_t1final.pdf. 

Rabalais, N.N., Turner, R.E., Sen Gupta, B.K., Platon, E., and Parsons, M.L., 2007, 

Sediments tell the history of eutrophication and hypoxia in the northern Gulf of 

Mexico: Ecological Applications, Supplement, v. 17, no. 5, p. 129-143. 

Ragauskas, A.J., Williams, C.K., Davison, B.H., Britovsek, G., Cairney, J., Eckert, C.A., 

Fredrick, W.J., Jr., Hallet, J.P., Leak, D.J., Liotta, C.L., Mielenz, J.R., Murphy, R., 
260

Templer, R., and Tschaplinski, T., 2006, The path forward for biofuels and 

biomaterials: Science, v. 311, p. 484–489. 

Ragueneau, O., Conley, D.J., Leynaert, A., Longphuirt, S.N., and Slomp, C.P., 2006a, 

Responses of coastal ecosystems to anthropogenic perturbations of silicon cycling, in

Ittekot, V., Unger, D., and Humborg, C., eds., The silicon cycle: Human 

perturbations and impacts on aquatic systems: Washington, D.C., Island Press, 

p. 197–213. 

Ragueneau, O., Conley, D.J., Leynaert, A., Longphuirt, S.N., and Slomp, C.P., 2006b, 

Role of diatoms in silica cycling and coastal marine food webs, in Ittekot, V., Unger, 

D., and Humborg, C., eds., The silicon cycle—Human perturbations and impacts on 

aquatic systems: Washington, D.C., Island Press, p. 163–195. 

Raloff, Janet, 2004, Dead waters—Massive oxygen-starved zones are developing along 

the world’s coasts: Science News Online, June 5, 2004, v. 165, no. 23, 8 p., accessed 

June 28, 2007, at http://sciencenews.org/articles/20040605/bob9.asp. 

Randall, G.W., Huggins, D.R., Russelle, M.P., Fuchs, D.J., Nelson, W.W., and Anderson, 

J.L., 1997, Nitrate losses through subsurface tile drainage in conservation reserve 

program, alfalfa, and row crop systems: Journal of Environmental Quality, v. 26, 

p. 1240–1247. 

Randall, G.W., and Sawyer, J., 2005, Nitrogen application timing, forms and additives, in

Gulf Hypoxia and Local Water Quality Concerns Workshop, Ames, IA, September 

27–28, 2005, p. 73–84. 

Randall, G.W., and Vetsch, J.A., 2005, Nitrate losses in subsurface drainage from a cornsoybean rotation as affected by fall vs. spring application of nitrogen and nitrapyrin: 

Journal of Environmental Quality, v. 34, no. 2, p. 590–597. 

Randall, G.W., Vetsch, J.A., and Huffman, J.R., 2003, Nitrate losses in subsurface 

drainage from a corn-soybean rotation as affected by time of nitrogen application 

and use of nitrapyrin: Journal of Environmental Quality, v. 32, p. 1764–1772. 

Raun, W.R., Solie, J.B., Stone, M.L., Martin, K.L., Freeman, K.W., Mullen, R.W., 

Zhang, H., Schepers, J.S., and Johnson, G.V., 2005, Optical sensor based algorithm 

for crop nitrogen fertilization: Communications in Soil Science and Plant Analysis, 

v. 36, nos. 19–20, p. 2759–2781. 

Redalje, D.G., Lohrenz, S.E., and Fahnenstiel, G.L., 1992, Phytoplankton dynamics and 

the vertical flux of organic carbon in the Mississippi River Plume and inner Gulf of 

Mexico shelf region, in Falkowski, P.G., and Woodhead, A.D., eds., Primary 

productivity and biogeochemical cycles in the sea: New York, Plenum Press, 526 p. 

Redalje, D.G., Lohrenz, S.E., and Fahnenstiel, G.L., 1994b, The relationship between 

primary production and the vertical export of particulate organic matter in a river 

impacted coastal ecosystem: Estuaries, v. 17, p. 829–838. 

Redalje, D.G., Lohrenz, S.E., and Fahnenstiel, G.L., 1994a, The vertical export of 

particulate and dissolved organic carbon from the surface waters of the northern Gulf 

of Mexico shelf: Journal of Mississippi Academy of Sciences, v. 39, p. 63. 
261

Reddy, K.R., Kadlec, R.H., Flag, E., and Gale, P.M., 1999, Phosphorous retention in 

streams and wetlands—A review: Environmental Science and Technology, Critical 

Review, v. 29, p. 83–146. 

Reddy, K.R., Wetzel, R.G., and Kadlec, R.H., 2005, Biogeochemistry of phosphorous in 

wetlands, in Phosphorous—Agriculture and the environment: Madison, WI, 

Agronomy Monograph no. 46, American Society of Agronomy, Crop Science 

Society of America, Soil Science Society of America, p. 263–316. 

Redfield, A.C., 1958, The biological control of chemical factors in the environment: 

American Scientist, v. 46, p. 205–222. 

Reetz, H.F., Jr., Murrell, T.S., and Murrell, L.J., 2001, Site-specific nutrient 

management—Production examples: Norcross, GA, Potash & Phosphate Institute, 

Better Crops, v. 85, no. 1, p. 12–13, 17. 

Reichelderfer, K., 1985, Do USDA program participants contribute to soil erosion: U.S. 

Department of Agriculture, Economic Research Service, Agricultural Economic 

Report No. 532, 83 p. 

Rhoads, D.C., Boesch, D.F., Tang, Z., Xu, F., Huang, L., and Nilsen, K.J., 1985, 

Macrobenthos and sedimentary facies on the Changjiang delta platform and adjacent 

continental shelf, East China Sea: Continental Shelf Research, v. 4, p. 189–213. 

Ribaudo, M., 1989, Targeting the Conservation Reserve Program to maximize water 

quality benefits: Land Economics, v. 65, p. 320–332. 

Ribaudo, M., Heimlich, R., Claassen, R., and Peters, M., 2001, Least-cost management of 

non-point source pollution—Source reduction versus interception strategies for 

controlling nitrogen loss in the Mississippi basin: Ecological Economics, v. 37, 

p. 183–197. 

Ribaudo, M., Heimlich, R., and Peters, M., 2005, Nitrogen sources and Gulf hypoxia—

Potential for environmental credit trading: Ecological Economics, v. 52, p. 159–168. 

Richards, R.P., and Baker, D.B., 2002, Trends in water quality in LEASEQ rivers and 

streams (northwestern Ohio), 1975–1995: Journal of Environmental Quality, v. 31, 

p. 90–96. 

Richards, R.P., Baker, D.B ., Kramer, J.W., Ewing, D.E., Merryfield, B.J., and Miller, 

N.L., 2001, Storm discharge, loads, and average concentrations in northwest Ohio 

rivers, 1975–1995: Journal of the American Water Resources Association, v.37, 

p. 423–438. 

Richardson, C.J., 1999, The role of wetlands in storage, release, and cycling of 

phosphorus on the landscape—A 25-year retrospective, in Reddy, K.R., ed., 

Phosphorus biogeochemistry in sub-tropical ecosystems, CRS Press/Lewis 

Publishers, Boca Raton, FL, p. 47–68. 

Richardson, W.B., Strauss, E.A., Bartsch, L.A., Monroe, E.M., Cavanaugh, J.C., 

Vingum, L., and Soballe, D.M., 2004, Denitrification in the upper Mississippi 

River—Rates, controls, and contribution to nitrate flux: Canadian Journal of 

Fisheries and Aquatic Science, v. 61, p. 1102–1112. 
262

Risser, P.G., 1985, Toward a holistic management perspective: BioScience, v. 35, 

p. 414–418.

Rizzo, W.M., Lackey, G.L., and Christian, R.R., 1992, Significance of euphotic, subtidal 

sediments to oxygen and nutrient cycling in a temperate estuary: Marine Ecology 

Progress Series, v. 86, p. 51–61. 

Robertson, W.D., Blowes, D.W., Placek, C.J., and Cherry, J.A., 2000, Long-term 

performance of in situ reactive barriers for nitrate remediation: Ground Water, v. 38, 

p. 689–695. 

Robinson, C.A., Cruse, R.M., and Ghaffarzadeh, M., 1996, Cropping system and nitrogen 

effects on mollisol organic carbon: Soil Science Society of America Journal, v. 60, 

p. 264–269. 

Rowe, G.T. and Chapman, P., 2002, Continental shelf hypoxia—Some nagging 

questions: Gulf of Mexico Science, v. 20, p. 155–160. 

Rowe, G.T., Cruz-Kaegi, M.L., Morse, J.W., Boland, G.S., and Escobar Briones, E.G., 

2002, Sediment community metabolism associated with continental shelf hypoxia, 

northern Gulf of Mexico: Estuaries, v. 25, no. 6, p. 1097–1106. 

Royer, T.V., David, M.B., and Gentry, L.E., 2006, Timing of riverine export of nitrate 

and phosphorus from agricultural watersheds in Illinois—Implications for reducing 

nutrient loading to the Mississippi River: Environmental Science and Technology, 

v. 40, p. 4126–4131. 

Royer, T.V., Tank, J.L., and David, M.B., 2004, The transport and fate of nitrate in 

headwater, agricultural streams in Illinois: Journal of Environmental Quality, v. 33, 

p. 1296–1304. 

Runkel, R.L., Crawford, C.G., and Cohn, T.A., 2004, Load estimator (LOADEST)—A 

FORTRAN program for estimating constituent loads in streams and rivers: U.S. 

Geological Survey Techniques and Methods, book 4, chap. A5, 69 p., available 

online at http://pubs.usgs.gov/tm/2005/tm4A5/. 

Russell, A.E., Laird, D.A., Parkin, T.B., and Mallarino, A.P., 2005, Impact of nitrogen 

fertilization and cropping system on carbon sequestration in Midwestern mollisols: 

Soil Science Society of America Journal, v. 69, p. 413–422. 

Russelle, M.P., Lamb, J.F.S., Montgomery, B.R., Elsenheimer, D.W., Miller, B.S., and 

Vance, C.P., 2001, Alfalfa rapidly remediates excess inorganic nitrogen at a fertilizer 

spill site: Journal of Environmental Quality, v. 30, p. 30–36. 

Santhi, C., Arnold, J.G., Williams, J.R., Dugas, W.A., Srinivasan, R., and Hauck, L.M., 

2001, Validation of the SWAT model on a large river basin with point and nonpoint 

sources: Journal of the American Water Resources Association, v. 37, no. 5, 

p. 1169–1188. 

Saunders, D.C., and Kalff, J., 2001, Nitrogen retention in wetlands, lakes and rivers: 

Hydrobiologia, v. 443, p. 205–212. 
263

Sawyer, J.E., and Nafziger, E.D., 2005, Regional approach to making nitrogen fertilizer 

rate decisions for corn: Proceedings of North Central Extension-Industry Soil 

Fertility Conference, v. 21, p. 16-24. 

Sawyer, J.E., and Randal, G.W., in press, Nitrogen rates, in Proceedings of Gulf Hypoxia 

and Local Water quality Concerns Workshop, Ames, IA, September 26-28, 2005. (in 

press)

Scavia, D., and Donnelly, K.A., in press, Reassessing hypoxia forecasts for the Gulf of 

Mexico: Environmental Science and Technology, (accepted for publication 

09/24/2007), available on line at: http://pubs.acs.org/cgibin/abstract.cgi/esthag/asap/abs/es0714235.html. 

Scavia, D., Justić, D., and Bierman, V.J., Jr., 2004, Reducing hypoxia in the Gulf of 

Mexico—Advice from three models: Estuaries, v. 27, no. 3, p. 419–425. 

Scavia, D., Kelly, E.L.A., and Hagy, J.D., III, 2006, A simple model for forecasting the 

effects of nitrogen loads on Chesapeake Bay hypoxia: Estuaries and Coasts, v. 29, 

no. 4, p. 674–684. 

Scavia, D., Rabalais, N.N., Turner, R.E., Justić, D., and Wiseman, W.J., Jr., 2003, 

Predicting the response of Gulf of Mexico hypoxia to variations in Mississippi River 

nitrogen load: Limnology and Oceanography, v. 48, no. 3, p. 951–956. 

Schaller, J.L, Royer, T.V., David, M.B., and Tank, J.L., 2004, Denitrification associated 

with plants and sediments in an agricultural stream: Journal of the North American 

Benthological Society, v. 23, no. 4, p. 667–676. 

Scharf, P.C., Brouder, S.M., and Hoeft, R.G., 2006a, Chlorophyll meter readings can 

predict nitrogen need and yield response of corn in the North-Central USA: 

Agronomy Journal, v. 98, p. 655–665. 

Scharf, P.C., Kitchen, N.R., Sudduth, K.A., Davis, J.G., 2006b, Spatially variable corn 

yield is a weak predictor of optimal nitrogen rate: Soil Science Society of America 

Journal, v. 70, p. 2154–2160. 

Scharf, P.C., Kitchen, N.R., Sudduth, K.A., Davis, J.G., Hubbard, V.C., and Lory, J.A., 

2005, Field-scale variability in optimal nitrogen fertilizer rate for corn: Agronomy 

Journal, v. 97, p. 452–461. 

Scheffer, M., Carpenter, S., Folley, J.A., Folke, C., and Walker, B., 2001, Catastrophic 

shifts in ecosystems: Nature, v. 413, p. 591–596. 

Schipper, L.A., Barkle, G.F., and Vojvodic-Vukovic, M., 2005, Maximum rates of nitrate 

removal in a denitrification wall: Journal of Environmental Quality, v. 34, p. 1270–

1276. 

Schipper, L.A., Barkle, G.F., Hadfield, J.C., Vojvodic-Vukovic, M., and Burgess, C.P., 

2004, Hydraulic constraints on the performance of a groundwater denitrification wall 

for nitrate removal from shallow groundwater: Journal of Contaminant Hydrology, 

v. 69, p. 263–279. 
264

Schipper, L., and Vojvodic-Vukovic, M., 1998, Nitrate removal from groundwater using 

a denitrification wall amended with sawdust—Field trial: Journal of Environmental 

Quality, v. 27, p. 664–668. 

Schipper, L.A., and Vojvodic-Vukovic, M., 2001, Five years of nitrate removal, 

denitrification and carbon dynamics in a denitrification wall: Water Research, v. 35, 

p. 3473–3477. 

Schlegel, A.J., Grant, C.A., and Havlin, J.L., 2005, Challenging approaches to nitrogen 

fertilizer recommendations in continuous cropping systems in the Great Plains: 

Agronomy Journal, v. 97, p. 391–398. 

Schlesinger, W.H., and Melack, J.M., 1981, Transport of organic carbon in the world’s 

rivers: Tellus, v. 33, p. 171–187. 

Schmidt, J.P., Schmitt, M.A., Randall, G.W., Lamb, J.A., Orf, J.H., and Gollany, H., 

2000, Swine manure application to nodulating and non-nodulating soybean: 

Agronomy Journal, v. 92, p. 987–992. 

Schmitt, M.A., Schmidt, D.R., and Jacobson, L.D., 1996, A manure management survey 

of Minnesota swine producers—Effect of farm size on manure application: Applied 

Engineering in Agriculture, v. 12, no. 5, p. 595–599. 

Schultz, R.C., Isenhart, T.M., Simpkins, W.W., and Colletti, J.P., 2004, Riparian forest 

buffers in agroecosystems—Lessons learned from the Bear Creek Watershed, central 

Iowa, USA: Agroforestry Systems, v. 61, p. 35–50. 

Scott, C.A., Walter, M.F., Brooks, E.S., Boll, J., Hes, M.B., and Merrill, M.D., 1998, 

Impacts of historical changes in land use and dairy herds in water quality in the 

Catskills Mountains: Journal of Environmental Quality, v. 27, p. 1410–1417. 

Segerson, K., and Miceli, T., 1998, Voluntary environmental agreements: Good or bad 

news for environmental protection?: Journal of Environmental Economics and 

Management, v. 36, p. 109–130. 

Seitzinger, S.P., and Giblin, A.E., 1996, Estimating denitrification in North Atlantic 

continental shelf sediments: Biogeochemistry, v. 35, p. 235–260. 

Seitzinger, S.P., Styles, R.V., Boyer, E.W., Alexander, R.B., Billen, G., Howarth, R.W., 

Mayer, B., and van Breemen, N., 2002, Nitrogen retention in rivers—Model 

development and application to watersheds in the northeastern U.S.A.: 

Biogeochemistry, v. 57–58, p. 199–237. 

Sferratore, A., Garnier, J., Billen, G., Conley, D.J., and Pinault, S., 2006, Diffuse and 

point sources of silica in the Seine River watershed: Environmental Science & 

Technology, v. 40, p. 6630–6635. 

Shapiro, C.A., Ferguson, R.B., Hergert, G.W., Dobermann, A.R., and Wortmann, C.S., 

2003, Fertilizer suggestions for corn: University of Nebraska-Lincoln Extension, 

G174, available on line at: http://elkhorn.unl.edu/epublic/live/g174/build/#target2. 

Sharpley, A.N., 1997, Rainfall frequency and nitrogen and phosphorus in runoff from soil 

amended with poultry litter: Journal of Environmental Quality, v. 26, p. 1127–1132. 
265

Sharpley, A.N., Daniel, T., Gibson, G., Bundy, L., Cabrera, M., Sims, T., Stevens, R., 

Lemunyon, J., Kleinman, P.J., and Parry, R., 2006a, Best management practices to 

minimize agricultural phosphorus impacts on water quality: U.S. Department of 

Agriculture, Agricultural Research Service ARS-163, 52 p. 

Sharpley, A.N., Lleinman, P.J.A., and McDowell, R.W., 2001, Innovative management 

of agricultural phosphorus to protect soil and water resources: Communications in 

Soil Science and Plant Analysis, v. 32, p. 1071–1100. 

Sharpley, A.N., Meisinger, J.J., Breeuwsma, A., Sims, J.T., Daniel, T.C., and Schepers, 

J.S., 1998, Impacts of animal manure management on ground and surface water 

quality, in Hatfield, J.L., and Stewart, B.A., eds., Animal waste utilization—

Effective use of manure as a soil resource: Boca Raton, FL, Ann Arbor Press, 

p. 173–242. 

Sharpley, A.N., Schmidt, J.P., and Hergert, L., 2006b, Nutrient management practices, in

Schnepf, M., and Cox, C., eds., Environmental benefits of conservation on 

cropland—The status of our knowledge: Ankeny, IA, Society of Soil and Water 

Conservation, p. 149–193. 

Sharpley, A.N., and Smith, S.J., 1994, Wheat tillage and water quality in the Southern 

Plains: Soil Tillage Research, v. 30, p. 33–38. 

Sharpley, A.N., Smith, S.J., and Bain, R., 1993, Effect of poultry litter application on the 

nitrogen and phosphorus content of Oklahoma soils: Soil Science Society of America 

Journal, v. 57, p. 1131–1137. 

Sharpley, A.N., Smith, S.J., Zollweg, J.A., and Coleman, G.A., 1996, Gully treatment 

and water quality in the Southern Plains: Journal of Soil and Water Conservation, 

v. 51, p. 512–517. 

Sharpley, A.N., Weld, J.L., Beegle, D.B., Kleinman, P.J.A., Gburek, W.J., Moore, P.A., 

Jr., and Mullins, G., 2003, Development of phosphorus indices for nutrient 

management planning strategies in the United States: Journal of Soil and Water 

Conservation, v. 58, no. 3, p. 137–152. 

Sharpley, A.N., Withers, P.J.A., Abdalla, C., and Dodd, A., 2005, Strategies for the 

sustainable management of phosphorus, in Sims, J.T., and Sharpley, A.N., eds., 

Phosphorus—Agriculture and the environment: American Society of Agronomy 

Monograph, p. 1069–1101. 

Shehandeh, H., Wright, A.L., Hons, F.M., and Lascano, R.J., 2005, Spatial and temporal 

variation of soil nitrogen parameters related to soil texture and corn yield: Agronomy 

Journal, v. 97, p. 772–782. 

Shepard, R., 2005, Nutrient management planning—Is it the answer to better 

management?: Journal of Soil and Water Conservation, v. 60, p. 171–176. 

Shibles, R.M., 1998, Soybean nitrogen acquisition and utilization, in Proceedings of the 

28th North Central Extension-Industry Soil Fertility Conference, St. Louis, MO, 11–

12 Nov. 1998: Brookings, SD, Potash & Phosphate Institute, p. 5–11. 
266

Shirmohammadi, A., Chaubey, I., Bosch, D.D., Muñoz-Carpena, R., Dharmasri, C., 

Arabi, M., Wolfe, M.L,. Frankenberger, J., Graff, C., Sohrabi, T.M., 

Shirmohammadi, A., 2006, Uncertainty in TMDL models: Transactions of the 

American Society of Agricultural and Biological Engineers, v. 49, no. 4, p. 1033–

1049. 

Simpson, J.H., and Hunter, J.R., 1974, Fronts in the Irish Sea: Nature, v. 250, p. 404–406. 

Sims, J.T., 1997, Agricultural and environmental issues in the management of poultry 

wastes— Recent innovations and long-term challenges, in Rechcigl, J.E., and 

MacKinnon, H.C., eds., Uses of by-products and wastes in agriculture: Washington, 

D.C., American Chemical Society, p. 72–90. 

Sims, J.T., Joern, B.C., and Simard, R.R., 1998, Phosphorus losses in agricultural 

drainage—Historical perspective and current research: Journal of Environmental 

Quality, v. 27, p. 277–293. 

Sims, J.T., and Kleinman, P.J.A., 2005, Managing agricultural phosphorus for 

environmental protection, in Sims, J.T., and Sharpley, A.N., eds., Phosphorus—

Agriculture and the environment: Madison, WI, American Society of Agronomy 

Monograph, American Society of Agronomy, p. 1021–1068. 

Singh, R., and Helmers, M.J., 2006, Subsurface drainage and its management in the 

upper Midwest tile landscape, in Proceedings of the 2006 EWRI Congress, American 

Society of Civil Engineers, Omaha, NE, May 21-25, 2006. 

Skaggs, R.W., Breve, M.A., and Gilliam, J.W., 1995, Predicting effects of water table 

management on loss of nitrogen from poorly drained soils: European Journal of 

Agronomy, v. 4, no. 4, p. 441–451. 

Skaggs, R.W., Youssef, M.A., and Chescheir, G.M., 2003, Effect of subsurface drain 

depth on nitrogen losses from drained lands: Transactions of the American Society 

of Agricultural Engineering, v. 46, p. 237–244. 

Skaggs, R.W., Youssef, M.A., Chescheir, G.M., Gilliam, J.W., 2005, Effect of drainage 

intensity on nitrogen losses from drained lands: Transactions of the American 

Society of Agricultural Engineering, v. 48, p. 2169–2177. 

Sklar, F.H., Chimney, M.J., Newman, S., McCormick, P., Gawlick, D., Miao, S.L., 

McVoy, C., Said, W., Newman, J., Coronado, C., Crozier, G., Korvela, M., and 

Rutchey, K., 2005, The ecological-societal underpinnings of Everglades restoration: 

Frontiers in Ecology and the Environment, v. 3, no. 3, p. 161–169. 

Sloth, N.P., Blackburn, H., Hansen, L.S., Risgaard-Petersen, N., and Lomstein, B.A., 

1995, Nitrogen cycling in sediments with different organic loading: Marine Ecology 

Progress Series, v. 116, p. 163–170. 

Smith, R.A., and Alexander, R.B., 2000, Sources of nutrients in the nation’s watersheds, 

in Managing nutrients and pathogens from animal agriculture, Proceedings from the 

Natural Resource, Agriculture, and Engineering Service Conference for Nutrient 

Management Consultants, Extension Educators, and Producer Advisors March 28–

30, 2000, Camp Hill, PA. 
267

Smith, R.A., Schwarz, G.E., and Alexander, R.B., 1997, Regional interpretation of water 

quality monitoring data: Water Resources Research, v. 33, p. 2781–2798. 

Smith, S.J., Sharpley, A.N., Berg, W.A., Naney, J.W., and Coleman, G.A., 1992, Water 

quality characteristics associated with Southern Plains grasslands: Journal of 

Environmental Quality, v. 21, p. 595–601. 

Smith, S.V., and Hollibaugh, J.T., 1989, Carbon-controlled nitrogen cycling in a marine 

‘macrocosm’—An ecosystem-scale model for managing cultural eutrophication: 

Marine Ecology Progress Series, v. 52, p. 103–109. 

Snyder, C.S., 2006, Phosphorus and potassium budgets and soil test levels in the 

Mississippi-Atchafalaya River Basin: Better Crops, v. 90, no. 1, p. 19–21, available 

online at 

http://www.ipni.net/ppiweb/bcrops.nsf/$webindex/D62FD5F4335D283E852571100

01556B9/$file/06-1p19.pdf. 

Snyder, C.S., Bruulsema, T.W., Sharpley, A.N., and Beegle, D.B., 1999, Site-specific use 

of the environmental phosphorus index tool: Norcross, GA, Potash & Phosphate 

Institute, SSMG-1, 4 p., available online at: http://www.ppippic.org/ppiweb/ppibase.nsf/b369c6dbe705dd13852568e3000de93d/1a2c31b028f94

9238525695300581e03/$FILE/SSMG1.pdf.

Snyder, C.S., and Leep, R.H., 2007, Fertilization, in Barnes, R.F., Nelson, C.J., Moore, 

K.J., and Collins, M., eds., Forages—Volume II, The science of grassland 

agriculture, 6th ed.: Ames, Iowa, Blackwell Publishing, chap. 24, p. 355–378. 

Snyder, C.S., Randall, G.W., Almond, R.E., and Hoeft, R.G., 2001, Fall nitrogen 

management for agronomic response and environmental protection: Fall fertilization 

facts—Opportunities and considerations: Norcross, GA, Potash & Phosphate 

Institute, available online at http://www.ppippic.org/ppiweb/ppibase.nsf/$webindex/BFA77B79E6C8028C8525694E002D096D

!opendocument

Sogbedji, M.J., and McIsaac, G.F., 2006, Evaluation of the ADAPT model for simulating 

nitrogen dynamics in a tile drained agricultural watershed in central Illinois: Journal 

of Environmental Quality, v. 35, p. 1914–1923. 

Sørensen, J., Rasmussen, L.K., and Koike, I., 1987, Micromolar sulfide concentrations 

alleviate acetylene blockage of nitrous oxide reduction by denitrifying Pseudomonas 

fluorescens: Canadian Journal of Microbiology, v. 33, p. 1001–1005. 

Spieles, D.J., and Mitsch, W.J., 2000, The effects of season and hydrologic and chemical 

loading on nitrate retention in constructed wetlands—A comparison of low and high 

nutrient riverine systems: Ecological Engineering, v. 14, p. 77–91. 

Sprague, L.A., Clark, M.L., Rus, D.L., Zelt, R.B., Flynn, J.L., and Davis, J.W., 2006, 

Nutrient and suspended-sediment trends in the Missouri River Basin, 1993–2003: 

U.S. Geological Survey Scientific Investigations Report 2006-5231, 80 p., available 

online at http://pubs.usgs.gov/sir/2006/5231/
268

St. John, J.P., Fitzpatrick, J.J., and Landeck Miller, R.E., in press, TMDL modeling for 

Long Island Sound: New York Water Environment Association, Clearwaters, v. 37, 

no. 3, (in press). 

Stadmark, J., and Leonardson, L., 2005, Emissions of greenhouse gases from ponds 

constructed for nitrogen removal: Ecological Engineering, v. 25, p. 542-551. 

Stevens, R.J., and Laughlin, R.J., 1998, Measurement of nitrous oxide and di-nitrogen 

emissions from agricultural soils: Nutrient Cycling in Agroecosystems, v. 52, p. 131-

139, available on-line at: http://www.springerlink.com/content/g61700k53q415214/. 

Stone, M.C., Hotchkiss, R.C., Hubbard, C.M., Fontaine, T.A., Mearnes, L.O., and 

Arnold, J.G., 2001, Impacts of climate change on Missouri River basin water yield: 

Journal of the American Water Resources Association, v. 37, no. 5, p. 1119–1130. 

Stow, C.A., Qian, S.S., and Craig, J.K., 2005, Declining threshold for hypoxia in the Gulf 

of Mexico: Environmental Science and Technology, v. 39, p. 716–723. 

Sugg, Zachary, 2007, Assessing U.S. farm drainage: Can GIS lead to better estimates of 

subsurface drainage extent?: Water Resources Institute, 8 p., available online at: 

http://www.wri.org/biodiv/pubs_description.cfm?pid=4324#pdf_files. 

Sutula, M., Bianchi, T.S., and McKee, B.A., 2004, Effect of seasonal sediment storage in 

the lower Mississippi River on the flux of reactive particulate phosphorus to the Gulf 

of Mexico: Limnology & Oceanography, v. 49, p. 2223–2235. 

Swaney, D.P., Shermana, D.M., and Howarth, R.W., 1996, Modeling water, sediment, 

and organic carbon discharges in the Hudson/Mohawk basin—Coupling two 

terrestrial sources: Estuaries, v. 19, no. 4, p. 833–847. 

Switzer, T.S., Chesney, E., and Baltz, D.M., 2006, Habitat selection by flatfishes along 

gradients of environmental variability—Implications for susceptibility to hypoxia in 

the northern Gulf of Mexico: Paper presented at Hypoxia Effects on Living 

Resources in the Gulf of Mexico, September 25–26, 2006, Tulane University, New 

Orleans, Louisiana, sponsored by National Oceanic and Atmospheric Administration 

Center for Sponsored Coastal Ocean Research. 

Sylvan, J.B., Dortch, Q., Nelson, D.M., Maier Brown, A.F., Morrison, W., and 

Ammerman, J.W., 2006, Phosphorus limits phytoplankton growth on the Louisiana 

shelf during the period of hypoxia formation: Environmental Science and 

Technology, v. 40, no. 24, p. 7548-7553, available on line at: 

http://pubs3.acs.org/acs/journals/supporting_information.page?in_manuscript=es061

417t. 

Taheripour, P., Khanna, M., and Nelson, C., 2005, Welfare impacts of alternative public 

policies for agricultural pollution control in an open economy: A general equilibrium 

framework: Presentation at the American Economic Association Annual Meeting, 

Province, Rhode Island, available online http://ageecon.lib.umn.edu/cgibin/pdf_view.pl?paperid=16472. 

Teisl, M., Roe, B., and Hicks, R., 2002, Can eco-labels tune a market?—Evidence from 

dolphin-safe labeling: Journal of Environmental Economics and Management, v. 43, 

p. 339–359. 
269

Terry, D., 2006, Fertilizer tonnage reporting in the U.S.—Basis and current need: Better 

Crops, v. 90, no. 4, p. 15–17, available online at 

http://www.ipni.net/ppiweb/bcrops.nsf/$webindex/F0BE4489F424A3FE8525721400

271DCB/$file/06-4p15.pdf.

Tetra Tech, Inc., 1998, Documentation of phase I and phase II activities in support of 

point source nutrient loading analysis in the Mississippi River system: Prepared for 

USEPA Nonpoint-source Control Branch, contract no. 68-C7-0014, Washington, 

D.C. 

Thogersen, J., 2002, Promoting ‘green’ consumer behavior with eco-labels, in Dietz, T., 

and Stern, P.C., eds., New Tools for Environmental Protection: Education, 

Information, and Voluntary Measures, The National Academies Press, National 

Academy of Sciences, p.83-104. 

Thom, W.O., and Sabbe, W.E., 1994, Soil sampling procedures for the southern region of 

the United States: Lexington, KY, Kentucky Agricultural Experiment Station, 

Southern Cooperative Series Bulletin 377. 

Thorpe, S.A., 2004, Langmuir circulation: Annual Review of Fluid Mechanics, v. 36, 

p. 55–79. 

Tiwari, A.K., Risse, L.M., and Nearing, M.A., 2000, Evaluation of WEEP and its 

comparison with USLE and RUSLE: Transactions of the American Society of 

Agricultural Engineers, v. 43, no. 5, p. 1129–1135. 

Tolbert, V., 1998, Guest editorial: Biomass & Bioenergy, v. 14, no. 4, p. 301–306. 

Tufekcioglu, A., Raich, J.W., Isenhart, T.M., and Schultz, R.C., 2003, Biomass, carbon 

and nitrogen dynamics of multi-species riparian buffers within an agricultural 

watershed in Iowa, USA: Agroforestry Systems, v. 57, p. 187–198. 

Turner, R.E., 1999, A comparative mass balance budget (C, N, P and suspended solids) 

for a natural swamp and overland flow systems, in Vymazal, J., ed., Nutrient cycling 

and retention in natural and constructed wetlands: Leiden, The Netherlands, 

Backhuys Publishing, p. 61–71. 

Turner, R.E., 2005, Nitrogen and phosphorus concentration and retention in water 

flowing over freshwater wetlands, in Fredrickson, L., King, S.L., and Kaminski, 

R.M., eds., Ecology and management of bottomland hardwood systems: Columbia, 

MO, The State of Our Understanding, University of Missouri Press, p. 57–66. 

Turner, R.E., Milan, C.S., and Rabalais, N.N., 2004, A retrospective analysis of trace 

metals, C, N and diatom remnants in the Mississippi River delta shelf: Marine 

Pollution Bulletin, v. 49, p. 548–556. 

Turner, R.E., Qureshi, N., Rabalais, N.N., Dortch, Q., Justic, D., Shaw, R.F., and Cope, 

J., 1998, Fluctuating silicate—Nitrate ratios and coastal plankton food webs: 

Proceedings of the National Academy of Sciences, v. 95, p. 13,048–13,051. 

Turner, R.E., and Rabalais, N.N., 1991, Changes in Mississippi River water quality this 

century—Implications for coastal food webs: BioScience, v. 41, p. 140–148. 
270

Turner, R.E., and Rabalais, N.N., 1994, Coastal eutrophication near the Mississippi River 

delta: Nature, v. 368, p. 619–621. 

Turner, R.E., Rabalais, N.N., Alexander, R.B., McIsaac, G., and Howarth, R.W., in press, 

Causes of Gulf of Mexico hypoxia 1—Characterization of nutrient and organic 

matter loads: Estuaries and Coasts, (in press). 

Turner, R.E., Rabalais, N.N., and Justić, D., 1999, Long-term watershed and water 

quality changes in the Mississippi River system, in Wiseman, W.J., Jr., Rabalais, 

N.N., Dagg, M.J., and Whitledge, T.E., eds., Nutrient enhanced coastal ocean 

productivity in the northern Gulf of Mexico: Silver Spring, Maryland, National 

Oceanic and Atmospheric Administration Coastal Ocean Program, Decision 

Analysis Series No. 14, U.S. Department of Commerce, National Ocean Service, 

Center for Sponsored Coastal Research, chap. 3, p. 37–50. 

Turner, R.E., Rabalais, N.N., and Justić, D., 2006, Predicting summer hypoxia in the 

northern Gulf of Mexico—Riverine N, P, and Si loading: Marine Pollution Bulletin, 

v. 52, p. 139–148. 

Twilley, R.R., Cowan, J., Miller-Way, T., Montagna, P.A., and Mortazavi, B., 1999, 

Benthic nutrient fluxes in selected estuaries in the Gulf of Mexico, in Bianchi, T.S., 

Pennock, J.R., and Twilley, R.R., eds., Biogeochemistry of Gulf of Mexico 

Estuaries: New York, Wiley, p. 163–209. 

Udawatta, R.P., Motavalli, P.P., and Garrett, H.E., 2004, Phosphorus loss and runoff 

characteristics in three adjacent agricultural watersheds with claypan soils: Journal of 

Environmental Quality, v. 33, p. 1709–1719. 

UMRSHNC, 2006, Gulf hypoxia and local water quality concerns workshop—A 

workshop assessing tools to reduce agricultural nutrient losses to water resources in 

the corn belt, in Workshop Proceedings, September 26–28, 2005, Iowa State 

University, Ames, Iowa, 205 p., available online at: 

http://www.umrshnc.org/index.php?option=com_content&task=view&id=19&Itemi

d=34

U.S. Department of Agriculture, 2003, Cost associated with development and 

implementation of Comprehensive Nutrient Management Plans—Part 1 – Nutrient 

management, land treatment, manure and wastewater handling and storage, and 

recordkeeping: U.S. Department of Agriculture, Natural Resource Conservation 

Service, 220 p., available online at 

http://www.nrcs.usda.gov/technical/land/pubs/cnmp1.html

U.S. Department of Energy, 2006, Clean coal and natural gas power systems: Available 

online at http://www.fe.doe.gov/programs/powersystems/index.html. 

U.S. Environmental Protection Agency, 2000a, State compendium—Programs and 

regulatory activities related to animal feeding operations: Washington, DC, USEPA 

Office of Water, Office of Waste Management. 

U.S. Environmental Protection Agency, 2000b, National air pollution trends, 1900–1998: 

Available online at: http://www.epa.gov/ttn/chief/trends/trends98/trends98.pdf. 
271

U.S. Environmental Protection Agency, 2000c, Ambient water quality criteria 

recommendations: Information supporting the development of state and tribal 

nutrient criteria: Rivers and streams in Ecoregion VI: Washington, DC, USEPA, 

Office of Water, EPA 822-B-00-017, 91 p., available online at: 

http://www.epa.gov/waterscience/criteria/nutrient/ecoregions/rivers/rivers_6.pdf. 

U.S. Environmental Protection Agency, 2000d, Nutrient criteria technical guidance 

manual: Rivers and Streams: Chapter 1 – Introduction: Washington, DC, USEPA 

Office of Water, Office of Science and Technology, EPA-822-B-00-002, 16 p., 

available online at: 

http://www.epa.gov/waterscience/criteria/nutrient/guidance/rivers/chapter_1.pdf. 

U.S. Environmental Protection Agency, 2002, National water quality inventory—2000 

Report to Congress: Washington, DC, USEPA, Office of Water, EPA-841-R-02-001, 

available online at: http://www.epa.gov/305b/2000report/. 

U.S. Environmental Protection Agency, 2003b, Economic analyses of nutrients and 

sediment reduction actions to restore Chesapeake Bay water quality: Annapolis, MD, 

EPA Chesapeake Bay Program Office, 162 p., available online at 

http://www.chesapeakebay.net/tribtools.htm

U.S. Environmental Protection Agency, 2003a, Nutrient reduction technology cost 

estimations for point sources in the Chesapeake Bay watershed: Annapolis, MD, 

EPA Chesapeake Bay Program Office, 132 p., available online at 

http://www.chesapeakebay.net/tribtools.htm

U.S. Environmental Protection Agency, 2004, Managing manure guidance for 

concentration animal feeding operations (CAFOs): Washington, D.C., USEPA, 

Office of Water, EPA-821-B-04-009, U.S. Government Printing Office, available at: 

http://cfpub.epa.gov/npdes/afo/info.cfm#manure

U.S. Environmental Protection Agency, 2004a, Water quality trading assessment 

handbook—Can water quality trading advance your watershed’s goals?: U.S. 

Environmental Protection Agency, Office of Water, Office of Wetlands, Oceans, and 

Watersheds, EPA 841-B-04-001, 120 p., available online at: 

http://www.epa.gov/owow/watershed/trading/handbook/docs/NationalWQTHandboo

k_FINAL.pdf. 

U.S. Environmental Protection Agency, 2004b, What is the status of point source 

nitrogen reduction in the Chesapeake Bay watershed?: Annapolis, MD, EPA 

Chesapeake Bay Program Office Fact Sheet, 2 p. 

U.S. Environmental Protection Agency, 2005, National management measures to control 

nonpoint source pollution from, urban areas: Washington, D.C., U.S. Environmental 

Protection Agency, Office of Water, EPA-841-B-05-004. 

U.S. Environmental Protection Agency, 2006b, Ecoregional nutrient criteria documents 

for rivers & streams: Available online at: 

http://www.epa.gov/waterscience/criteria/nutrient/ecoregions/rivers/. 
272

U.S. Environmental Protection Agency, 2006a, National Emissions Inventory (NEI) Air 

Pollutant Emissions Trends Data: Available online at 

http://www.epa.gov/ttn/chief/trends/index.html. 

U.S. Environmental Protection Agency, 2007, The Long Island Sound study at EPA New 

England: available on line at http://www.epa.gov/boston/eco/lis/epane.html. 

U.S. Government Accounting Office, 1997, Global warming—Information on the results 

of four of EPA’s voluntary climate change programs: Washington, D.C., U.S. 

Government Printing Office, GAO:RCED-97-163, 32 p. Available on line at: 

http://www.gao.gov/archive/1997/rc97163.pdf.

Vache, K.B., Eilers, J.M., and Santelman, M.V., 2002, Water quality modeling of 

alternative agricultural scenarios in the U.S. corn belt: Journal of the American 

Water Resources Association, v. 38, no. 2, p. 773–787. 

Vahtera, E., Conley, D., Gustufsson, B.G., Kuosa, H., Pitkanen, H., Savchuk, O.P., 

Tamminen, T., Vitasalo, M., Voss, M., Wasmund, N., and Wulff, F., 2007, Internal 

ecosystem feedbacks enhance nitrogen-fixing cyanobacteria blooms and complicate 

management in the Baltic Sea: Ambio, v. 36, no. 12, p. 186–194. 

Van Driel, P.W., Merkley, L.C., and Robertson, W.D., 2006, Denitrification of 

agricultural drainage using wood-based reactors: Transactions of the ASABE, v. 49, 

p. 565–573. 

Van Liew, M.W., Veith, T.L., Bosch, D.D., and Arnold, J.G., 2006, Suitability of SWAT 

for the Conservation Effects Assessment Project—A comparison on USDA ARS 

watersheds: Journal of Hydrologic Engineering, v.12, no. 2, p. 173-183. 

Varvel, G.E., 2006, Soil organic carbon changes in diversified rotations of the western 

cornbelt: Soil Science Society of America Journal, v. 70, p. 426–433. 

Varvel, G.E., Schepers, J.S., and Francis, D.D., 1997, Ability for in-season correction of 

nitrogen deficiency in corn using chlorophyll meters: Soil Science Society of 

America Journal, v. 61, p. 1233–1239. 

Vollenweider, R.A., 1976, Advances in defining critical loading levels of phosphorus in 

lake eutrophication: Memorie dell' Istituto Italiano di Idrobiologia, v. 33, p. 53–83. 

Von Holle, C.K., 2005, Agricultural nitrogen use and producer attitudes in tile-drained 

watersheds of east-central Illinois: University of Illinois at Urbana-Champaign M.S. 

Thesis, Urbana, IL, 72 p. 

Walker, N.D., and Rabalais, N.N., 2006, Relationships among satellite chlorophyll a, 

river inputs, and hypoxia on the Louisiana continental shelf, Gulf of Mexico: 

Estuaries and Coasts, v. 29, no. 6B, p. 1081–1093. 

Wang, X., and Melesse, A.M., 2005, Evaluation of the SWAT model’s snowmelt 

hydrology in a northwestern Minnesota watershed: Transactions of the American 

Society of Agricultural and Biological Engineers, v. 48, no. 4, p. 1359–1376. 

Watershed Agriculture Council, 2004, History of the Watershed Agriculture Council: 

Available online at http://www.nycwatershed.org/index_wachistory.html
273

Wawrik, B., Paul, J.H., Bronk, D.A., and Gray, J.D., 2004, High rates of ammonium 

recycling drive phytoplankton productivity in the offshore Mississippi River plume: 

Aquatic Microbial Ecology, v. 35, p. 175–184. 

Wedwick, S., Lakhani, B., Stone, J., Waller, P., and Artiola, J., 2001, Development and 

sensitivity analysis of the GLEAMS-IR model: Transactions of the American 

Society of Agricultural Engineers, v. 44, no. 5, p. 1095–1104. 

Weed, D.A.J., and Kanwar, R.S., 1996, Nitrate and water present in and flowing from 

root zone soil: Journal of Environmental Quality, v. 25, p. 709-719. 

Wells, M.C., Ju, Z., Heater, S.J., and Walter, R.B., 2006, Microarray gene expression 

analyses in Medaka (Oryzias latipes) exposed to hypoxia: Paper presented at 

Hypoxia Effects on Living Resources in the Gulf of Mexico, September 25–26, 2006 

Tulane University, New Orleans, Louisiana, sponsored by National Oceanic and 

Atmospheric Administration Center for Sponsored Coastal Ocean Research. 

Westerman, P.W., Donnely, T.L., and Overcash, M.R., 1983, Erosion of soil and poultry 

manure—A laboratory study: Transactions of the American Society of Agricultural 

Engineers, v. 26, p. 1070–1078, 1084. 

Wilhelm, W.W., Johnson, J.M.F., Hatfield, J.L., Voorhees, W.B., and Linden, D.R., 

2004, Crop and soil productivity response to corn residue removal—A literature 

review: Agronomy Journal, v. 96, no. 1, p. 1–17. 

Wiseman, W.J., Rabalais, N.N., Turner, R.E., Dinnel, S.P., and MacNaughton, A., 1997, 

Seasonal and interannual variability within the Louisiana coastal current: 

stratification and hypoxia: Journal of Marine Systems, v.12, p. 237–248. 

Wiseman, Jr., W.J., Rabalais, N.N., Turner, R.E., and Justic, D., 2004, Hypoxia and the 

physics of the Louisiana coastal current, in Nihoul, J.C.J., Zavialov, P.O., and 

Micklin, P.P., eds., Dying and Dead Seas – Climate Versus Anthropogenic Causes, 

Kluwer Academic Publishers, Dordrecht, p. 359–372. 

Wisner, Robert, 2007, Iowa Farm Outlook: Presentation at Tel Aviv University, May 15, 

2007, available online at: 

http://www.econ.iastate.edu/facultywisner/documents/telavivethanolpresentationwisner07.pdf

Wittry, D.J., and Mallarino, A.P., 2002, Use of variable-rate technology for agronomic 

and environmental phosphorus-based liquid swine manure management, in Robert, 

P.C., Rust, R.H., and Larson, W.E., eds., Sixth International Conference on SiteSpecific Management for Agricultural Systems Proceedings, American Society of 

Agronomy, July 14–17, 2002, Minneapolis, MN. 

Wittry, D.J., and Mallarino, A.P., 2004, Comparison of uniform- and variable-rate 

phosphorus fertilization for corn-soybean rotations: Agronomics Journal, v. 96, 

p. 26–33. 

Wollheim, W.M., Vorosmarty, C.J., Peterson, B.J., Seitzinger, S.P., and Hopkinson, C.S., 

2006, Relationship between river size and nutrient removal: Geophysical Research 

Letters, v. 33, (L06410): doi:10.1029/2006GL025845. 
274

Wortmann, C., Helmers, M., Mallarino, A.P., Barden, C., Devlin, D., Pierzynski, G., 

Lory, J., Massey, R., Holz, J., Shapiro, C., and Kovar, J., 2005, Agricultural 

phosphorus management and water quality protection in the Midwest: Lincoln, NE, 

University of Nebraska-Lincoln Extension and CSREES-USDA, Heartland Regional 

Water Coordination Initiative, Regional Publication 187, 24 p., available online at: 

http://www.ianrpubs.unl.edu/epublic/live/rp187/build/rp187.pdf.

Wu, J., Adams, R., Kling, C., and Tanaka, K., 2004, Assessing the costs and 

environmental consequences of agricultural land use changes—A site-specific, 

policy-scale modeling approach: American Journal of Agricultural Economics, v. 86, 

p. 26–41. 

Wu, J., and Babcock, B., 1999, Metamodeling potential nitrate water pollution in the 

central United States: Journal of Environmental Economics and Management, v. 28, 

p. 1916–1928. 

Wu, J., and Tanaka, K., 2005, Reducing nitrogen runoff from the Upper Mississippi 

River basin to control hypoxia in the Gulf of Mexico—Easements or taxes?: Marine 

Resource Economics, v. 20, p. 121–144. 

Wysocki, L.A, Bianchi, T.S., Powell, R., and Reuss, N., 2006, Spatial variability in the 

coupling of organic carbon, nutrients, and phytoplankton pigments in surface waters 

and sediments of the Mississippi River plume: Estuarine, Coastal and Shelf Science, 

v. 69, p. 47–63. 

Xu, Y. Jun, 2006, Total nitrogen inflow and outflow from a large river swamp basin to 

the Gulf of Mexico: Hydrological Sciences – Journal – des Sciences Hydrologiques, 

v. 51, no. 3, p. 531–542. 

Xue, Y., David, M.B., Gentry, L.E., and Kovacic, D.A., 1998, Kinetics and modeling of 

dissolved phosphorus export from a tile-drained agricultural watershed: Journal of 

Environmental Quality, v. 27, p. 917–922. 

Yang, W., Khanna, M., and Farnsworth, R., 2005, Effectiveness of conservation 

programs in Illinois and gains from targeting: American Journal of Agricultural 

Economics, v. 5, p. 1248–1255. 

Yang, W., Khanna, M., Farnsworth, R., and Onal, H., 2004, Is geographical targeting 

cost-effective: The case of the Conservation Reserve Enhancement Program in 

Illinois: Review of Agricultural Economics, v. 27, p. 70-88. 

Yang, W., Khanna, M., Farnsworth, R., and Onal, H., 2003: Integrating economics, 

environmental and GIS modeling to target cost effective land retirement in multiple 

watersheds: Ecological Economics, v. 46, p. 249-267. 

Yuan, Y., Bingner, R.L., and Rebich, R.A., 2001, Evaluation of AnnAGNPS on 

Mississippi Delta MSEA watersheds: Transactions of the ASAE, v. 44, no. 5, 

p. 1183–1190. 

Yuan, T., Bingner, R.L., Theurer, F.D., 2006, Subsurface flow component for 

AnnAGNPS: Applied Engineering in Agriculture, v. 22, no. 2, p. 231–241. 
275

Yuan, Y., Bingner, F.D., Theurer, F.D., Ribich, R.A., and Moore, P.A., 2005, Phosphorus 

component in AnnAGNPS: Transactions of the American Society of Agricultural 

Engineers, v. 48, no. 6, p. 2145–2154. 

Zavala-Hidalgo, J., Morey, S.L., and O’Brien, J.J., 2003, Seasonal circulation on the 

western shelf of the Gulf of Mexico using a high-resolution numerical model: 

Journal of Geophysical Research, v. 108, C12, 3389, doi:10.1029/2003JC001879. 

Zhang, L., and Mitsch, W.J., 2000, Hydrologic budgets of the two Olentangy River 

experimental wetlands, 1994–99, in Mitsch, W.J., and Zhang, L., eds., Olentangy 

River Wetland Research Park at the Ohio State University, Annual Report 1999, 

p. 41–46. 

Zhang, L., and Mitsch, W.J., 2001, Water budgets of the two Olentangy River 

experimental wetlands in 2000, in Mitsch, W.J., and Zhang, L., eds., Olentangy 

River Wetland Research Park at the Ohio State University, Annual Report 2000, 

p. 17–28. 

Zhang, L., and Mitsch, W.J., 2002, Water budgets of the two Olentangy River 

experimental wetlands in 2001, in Mitsch, W.J., and Zhang, L., eds., Olentangy 

River Wetland Research Park at the Ohio State University, Annual Report 2001, 

p. 23–34. 

Zhang, L., and Mitsch, W.J., 2004, Water budgets of the two Olentangy River 

experimental wetlands in 2003, in Mitsch, W.J., Zhang, L., and Tuttle, C., eds., 

Olentangy River Wetland Research Park at the Ohio State University, Annual Report 

2003, p. 39–52. 

Zimmerman, A.R., and Nance, J.M. 2001, Effects of hypoxia on the shrimp industry of 

Louisiana and Texas, in Rabalais, N.N., and Turner, R.E., eds., Coastal hypoxia—

Consequences for living resources coastal and estuarine studies: Washington, D.C., 

American Geophysical Uniton, chap. 15, v. 58, p. 293–310. 

Zou, E., 2006, Impacts of hypoxia on physiology and toxicology of the brown shrimp 

Penaeus aztecus: Paper presented at Hypoxia Effects on Living Resources in the 

Gulf of Mexico, September 25-26, 2006, Tulane University, new Orleans, Louisiana, 

sponsored by National Oceanic and Atmospheric Administration Center for 

Sponsored Coastal Ocean Research. 
Appendix A 

A-1

Appendices 

A. Appendix A: Studies on the Effects of Hypoxia on Living Resources 

The abstracts in this appendix all came from a workshop sponsored by the NOAA 

Center for Sponsored Coastal Ocean Research, held at Tulane University, New Orleans, 

LA held September 25-26, 2006. 

Brouwer, Marius, 2006. “Changes in Gene and Protein Expression and Reproduction in 

Grass Shrimp, Palaemonetes pugio, Exposed to Chronic Hypoxia” Presentation at 

“Hypoxia Effects on Living Resource in the Gulf of Mexico” NOAA Center for 

Sponsored Coastal Ocean Research, Tulane University, New Orleans, LA. September 25 

– 26, 2006. 

Abstract: Hypoxic conditions in estuaries are one of the major factors responsible for declines in 

habitat quality. Previous studies examining the effects of hypoxia on crustacea have focused on 

individual/population-level, physiological or molecular responses but have not considered more 

than one type of response in the same study. The objective of this study was to integrate 

disciplines by examining the responses of grass shrimp to chronic hypoxia both at the molecular 

and whole animal level. Hypoxia-induced alterations in gene expression were screened using 

custom cDNA macroarrays containing 78 clones from a hypoxia-responsive suppression 

subtractive hybridization (SSH) cDNA library. Grass shrimp respond differently to moderate (2.5 

ppm DO) versus severe (1.5 ppm DO) chronic hypoxia. The initial response to moderate hypoxia 

was down-regulation of genes coding for ribosomal proteins, HSP 70 and MnSOD. The initial 

response after short-term (3 d) exposure to severe hypoxia was upregulation of genes involved in 

oxygen uptake/transport and energy production, such as hemocyanin and ATP synthases. The 

major response by day 7 was an increase of transcription of genes present in the mitochondrial 

genome, together with upregulation of a putative heme binding protein and the iron storage 

protein, ferritin. By day 14 a dramatic reversal was seen, with a significant downregulation of 

transcription of genes in the mitochondrial genome. Both ferritin and the heme binding protein 

were downregulated as well. Levels of Hypoxia Inducible Factor (HIF1-alpha) remained 

unchanged. The macroarray data were validated using real-time qPCR. Changes in mitochondrial 

proteins were examined by separating proteins in 2 dimensions (IEF and reverse phase) followed 

by MS. At the organismal level, hypoxia exposure resulted in marked effects on shrimp egg 

production and larval survival, suggesting population-level implications of long-term hypoxia. 

Baltz, Donald M., Hiram W. Li, Philippe A. Rossignol, Edward J. Chesney and 

Theodore S. Switzer, 2006. “A Qualitative Assessment of the Relative Effects of 

Bycatch Reduction,Fisheries and Hypoxia on Coastal Nekton Communities in the Gulf of 

Mexico”, Presentation at “Hypoxia Effects on Living Resource in the Gulf of Mexico” 

NOAA Center for Sponsored Coastal Ocean Research, Tulane University, New Orleans, 

LA. September 25 – 26, 2006. 

Abstract: We applied qualitative mathematical models to develop an understanding of linkages 

that influence shrimp, fishes, and fisheries in coastal Louisiana where biotic communities face 

many natural and anthropogenic stressors, one of which is fishing activities related to the harvest 

of shrimp. Shrimp trawling ranks high in terms of impact on nekton and their habitats, and like 

most fishing gears catches non-target species or sizes that are not marketed. These individuals, 

termed 'bycatch', are often returned to the water in dead or dying condition. Numerous other 
Appendix A 

A-2

individuals are not 'caught' per se but also suffer the 'effects of fishing', that can degrade habitats 

or cause injuries leading to mortality. Modeling was used to examine the effects of fishing and 

bycatch mortality on community structure in the 'Fertile Fisheries Crescent' and how major 

stressors interact with hypoxia to influence fisheries. We explored direct and indirect interactions 

between shrimp, their predators, bycatch species, and shrimp landings. A major finding was that 

bycatch reduction efforts may feedback on fisheries and shrimp populations in an unexpectedly 

negative manner. Another was that changes in community structure that might be attributed to 

hypoxia are also possible from fishing alone. To corroborate our models, we analyzed 15 years of 

quantitative data on National Marine Fisheries Service shrimp landings, Louisiana Department of 

Wildlife and Fisheries (LDWF) gillnet surveys, and LDWF shrimp trawl surveys from central 

Louisiana. Abundant bycatch and other species were summarized into several functional groups 

including small and large shrimp predators, non-shrimp predators, major bycatch consumers, 

minor bycatch consumers, and non-bycatch consumers. Factor and correlation analyses of 

quantitative data for functional groups on a bimonthly basis corroborated results from the 

qualitative models, and combined indicated that shrimp abundance and shrimp landings would 

likely suffer from increased natural mortality if the shrimp-fishery bycatch was substantially 

reduced. 

Craig, J. Kevin and Larry B. Crowder, 2005. “Hypoxia-induced habitat shifts and 

energetic consequences in Atlantic croaker and brown shrimp on the Gulf of Mexico 

shelf” Marine Ecology Progress Series, Vol. 294, pp 79-94. 

Abstract: This paper evaluates the effects of hypoxia-induced habitat loss on Atlantic croaker and 

brown shrimp. The compare spatial distributions and the relationship to abiotic factors, including 

temperature, dissolved oxygen and salinity across years with differing levels of hypoxia using 14 

years of fishery-independent trawl data. They find that hypoxia results in considerable shifts in 

temperature and oxygen conditions that croaker and brown shrimp experience. Croaker typically 

occupy relative warm, inshore waters. During periods of hypoxia, croaker remain in the warmest 

inshore waters, but are also displaced to cooler offshore waters. Brown shrimp typically are 

distributed more broadly and further offshore. During periods of hypoxia, brown shrimp shift to 

warm inshore waters and cooler waters near the offshore edge of the hypoxic zone. The shifts in 

spatial distribution are reflected in decreases in water temperature for croaker that are displaced 

offshore the hypoxic region, and increases in water temperature for brown shrimp that are displace 

inshore of the hypoxic zone. Both species also face increased variance in water temperatures due 

to hypoxia-induced habitat displacement. Despite avoidance of the lowest oxygen waters, high 

densities of croaker and brown shrimp occur in areas of 1.6 to 3.7 mg/l near the offshore hypoxic 

edge. Shifts in spatial distribution during severe hypoxia may impact organism energy budgets. 

For example, laboratory studies indicate low oxygen impacts individual movement, growth, and 

mortality (Wannamaker & Rice 2000, Taylor & Miller 2001, Wu 2002). High croaker and shrimp 

densities near the hypoxic edge likely have implications for trophic interactions as well as the 

harvest of both target (brown shrimp) and nontarget (croaker) species by the commercial shrimp 

fishery. Croaker may benefit from high concentrations of brown shrimp at the edge of the hypoxic 

zone, while brown shrimp may become more susceptible to predation by croaker. 

Craig, J. Kevin, Larry B. Crowder, and Tyrrell A. Henwood, 2005. “Spatial distribution 

of brown shrimp (Farfantepenaeus aztecus) on the northwestern Gulf of Mexico shelf: 

effects of abundance and hypoxia” Canadian Journal of Fisheries and Aquatic Science. 

Vol. 62 pp 1295-1308. 

Abstract: This paper uses fishery-independent hydrographic and bottom trawl surveys from 1983–

2000 used to test for density dependence and effects of hypoxia on spatial distribution of brown 

shrimp. The spatial distribution of shrimp was found to be positively related to abundance on the 

Texas shelf, but negatively related to abundance on the Louisiana shelf. Density dependence was 

weak, and may have been due to factors other than habitat selection. Large-scale hypoxia (up to 
Appendix A 

A-3

~20 000 km2) on the Louisiana shelf occurs in regions of typically high shrimp density, resulting 

in loss of up to 25% of shrimp habitat on the Louisiana shelf. They also find shifts in distribution 

and densities both inshore and offshore of the hypoxic region. Results placed in terms of the 

generality of density-dependent spatial distributions in marine populations. Potential consequences 

of habitat loss and associated shifts in distribution due to low dissolved oxygen. They note that 

shifts in spatial distribution may precede major stock declines, and thus could potentially serve as 

an early warning sign of future declines in abundance (Gomes et al.1995; Rose et al. 2000; 

Overholtz 2002). 

Diaz, Robert, 2001. “Overview of Hypoxia around the World” Journal of Environmental 

Quality Vol. 30, No. 2. (March-April) 275-281. 

Abstract: This paper summarizes effects of hypoxia in various locations around the world, which 

provides lessons for potential consequences of hypoxia in the Gulf of Mexico. They note that 

hypoxia was probably not a prominent feature of the shallow continental shelf in the Northern 

Gulf of Mexico prior to the 1920’s through 1950’s based on geo-chronology of sediment cores. A 

longer, 2000-year chronology in the Chesapeake indicates that early European settlement of the 

watershed was a key feature that set the stage for current oxygen problems. Improved water 

quality in Lake Erie is the best example in the US that large ecosystems can respond positively to 

nutrient regulation, but the time interval for recovery can be long. In Lake Erie, the extent of 

hypoxia was similar between 1970 and 1990 despite reduced nutrient loads. Delayed 

improvements in oxygen levels are argued to be consistent with mechanisms and processes that 

contribute to ecosystem’s resilience (Charlton et al, 1993), and as a consequence improvements in 

oxygen may not be noticed for decades following implementation of management actions. 

Hendon, Laura A. Erik A. Carlson, Steve Manning, and Marius Brouwer, 2006. “Crosstalk between Pyrene and Hypoxia Signaling Pathways in Embryonic Cyprinodon 

variegates” Presentation at “Hypoxia Effects on Living Resource in the Gulf of Mexico” 

NOAA Center for Sponsored Coastal Ocean Research, Tulane University, New Orleans, 

LA. September 25 – 26, 2006. 

Abstract: The aryl hydrocarbon nuclear translocator (ARNT) is a general dimeric partner for the 

aryl hydrocarbon receptor (AhR) and hypoxia-inducible factor one alpha (HIF1-α). The 

AhR/ARNT complex binds to promoters in target genes, such as CYP1A1, resulting in alterations 

in gene expression, while the HIF1-α/ARNT heterodimer binds to hypoxia response elements in 

target genes, such as VEGF. While AhR is activated by PAHs, such as pyrene, HIF1-α is 

activated by hypoxia. Since ARNT is a general dimeric partner for both AhR and HIF1-α, 

possible cross-talk may exist between the two pathways in which the activation of one results in 

inhibition of the other. The objective of this study was to determine if pyrene-activation of AhR2, 

or hypoxia-activation of HIF1-α could sequester the ARNT protein away from HIF1-α and AhR2, 

respectively, resulting in reduced developmental toxicity associated with hypoxia or pyrene alone 

in embryonic Cyprinodon variegatus. As a first step to examine this hypothesis, we cloned AhR2, 

CYP1A1 (PAH-activated gene) and VEGF (HIF-activated gene). Next, pyrene (20, 60, and 150 

ppb) and hypoxia’s (1-2 ppm) individual developmental toxicity endpoints were determined, 

together with CYP1A1 and VEGF expression levels using real-time quantitative RT-PCR. 

Combined treatments of pyrene and hypoxia were examined in order to determine sequestration of 

the ARNT protein and developmental toxicity endpoints. Results demonstrate that pyrene-treated 

embryos alone develop toxicity endpoints such as pericardial edema and dorsal body curvature. 

Hypoxia-treated embryos alone display delayed hatching and less-developed characteristics in 

comparison to normoxic treatments. Under hypoxic conditions alone, real-time quantitative RTPCR determined that VEGF was down-regulated significantly at 24 hpf, while at 14 dph, the HIFactivated gene was significantly up-regulated. Pyrene-treated embryos showed a dose-dependent 

and time-dependent response in CYP1A1 regulation with increasing expression over time of 
Appendix A 

A-4

exposure. The combined effects of pyrene and hypoxia appeared to alter VEGF expression, while 

CYP1A1 remained unaffected in C. variegatus. 

Montagna, Paul, Ben Hodges, David Maidment and Barbara Minsker, 2006. “LongTerm Studies of Hypoxia in Corpus Christi Bay: The Cybercollaboratory Testbed” 

Presentation at “Hypoxia Effects on Living Resource in the Gulf of Mexico” NOAA 

Center for Sponsored Coastal Ocean Research, Tulane University, New Orleans, LA. 

September 25 – 26, 2006. 

Abstract: Corpus Christi Bay is a shallow (~3.2 m) enclosed bay with a level bottom. It 

experiences high wind speeds, temperatures, and receives a low amount of fresh water inflow. 

Hypoxia has been documented in the southeastern region of Corpus Christi Bay every summer 

since 1988. Hypoxia found in bottom waters, usually within 1 m from bottom, when the bay is 

stratified. Over the last 20 years, there has been increased surface water temperatures, but no 

change in nutrient concentrations, which are low. Ecosystem processes during salinity 

stratification likely drive the hypoxia, because respiration is stimulated and the surface and bottom 

water masses are not mixing. Hypoxia causes reduced benthos abundance, biomass, and diversity. 

The reduction is due to loss of deeper-dwelling organisms, and is likely a direct effect (stress or 

death), and not an indirect effect (increased predation by exposure to the surface). There is 

increased interest in developing real-time environmental forecasting and management to better 

monitor and understand large-scale, event-based environmental phenomena, e.g., hypoxia and 

flooding. A new project focuses on creating a new Corpus Christi Bay Observatory Testbed 

Project to demonstrate how cyberinfrastructure can enable real-time forecasting from a 

hydrographic information system. Although only a few months old, the testbed project has 

already created a few simple models and visualization tools that improved sampling designs to 

better identify hypoxic events, extent, and intensity. 

O’Connor, Thomas and David Whitall, 2007. “Linking Hypoxia to Shrimp Catch in the 

Northern Gulf of Mexico”, Marine Pollution Bulletin Vol. 54, no. 4 (April), Pp 460-463. 

Abstract: This study carries out updates the statistical analysis of Zimmerman and Nance () of the 

effect of hypoxia on commercial shrimp landings data for 1985 through 2004. This study uses 

commercial landings data, not the interview data, and is therefore does not use spatial data on the 

location of catch. The paper confirms the results of Zimmerman and Nance that there is no 

correlation of hypoxic area with landings of white shrimp or with landings of brown shrimp in 

Louisiana, but there is a significant correlation with the total combined landings in Texas and 

Louisiana. Unlike Zimmermann and Nance, they find a significant relationship between the 

hypoxic area and brown shrimp landings in Texas alone. Hypoxia explains about 32% of the 

variance in catch using data for catch in July and August, and about 27% of the variance in catch 

using annual data. 

Perez, Amy N., Leon Oehlers and Ronald B. Walter, 2006. “Detection of Hypoxiarelated Proteins in Medaka (Oryzias latipes) by Difference Gel Electrophoresis and 

Identification by Sequencing of Peptides using MALDI-TOF Mass Spectrometry” 

Presentation at “Hypoxia Effects on Living Resource in the Gulf of Mexico” NOAA 

Center for Sponsored Coastal Ocean Research, Tulane University, New Orleans, LA. 

September 25 – 26, 2006. 

Abstract: Multidimensional separation techniques combined with matrix-assisted laser 

desorption/ionization tandem time-of-flight mass spectrometry (MALDI-TOF/TOF-MS) were 

used to identify hypoxia-related biomarker proteins in tissues of medaka fish (Oryzias latipes) and 

medaka cultured cells. The multidimensional protein/peptide separation methods used included 
Appendix A 

A-5

two-dimensional difference gel electrophoresis (2D-DIGE) using fluorescent cyanine dyes, and 

gel electrophoresis combined with reversed phase liquid chromatrography of tryptic peptides 

isotopically labeled with 16O or 18O (geLC-MS). In both methods, control and hypoxia-treated 

tissue or cell protein extracts were differentially labeled, combined in 1:1 mass ratios, and 

subjected to separation and MALDI-TOF/TOF-MS analysis of tryptic peptides derived from 

proteins exhibiting significant changes in expression upon hypoxia exposure. Prior to MALDITOF/TOF-MS analysis, the peptides were N-terminally sulfonated using the derivatizing reagent 

4-sulfophenyl isothiocyanate (SPITC) to enhance the post-source decay (PSD) fragmentation 

spectra of the peptides in MALDI-TOF/TOF-MS, which was shown to dramatically improve de 

novo sequencing of labeled peptides. The methods described here were used to monitor and 

analyze the changes in protein resulting from exposures of both cultured medaka cells and medaka 

fish to hypoxic conditions (0.8-1.0 mg/L dissolved oxygen) for periods up to 120 hours. We have 

identified a number of potential candidate biomarker proteins differentially-regulated upon 

exposure to hypoxia, including carbonic anhydrase, hemoglobin, calbindin, aldolase, glutathioneS-transferase, succinate dehydrogenase, and lactate dehydrogenase. 

Rabalais, Nancy N. 2006. “Benthic Communities and the Effects of Hypoxia in 

Louisiana Coastal Waters” Presentation at “Hypoxia Effects on Living Resource in the 

Gulf of Mexico” NOAA Center for Sponsored Coastal Ocean Research, Tulane 

University, New Orleans, LA. September 25 – 26, 2006. 

Abstract: The responses of the benthic fauna to decreasing concentration of dissolved oxygen 

follow a fairly consistent pattern of progressive stress and mortality as the oxygen concentration 

decreases from 2 mg l-1 to anoxia (0 mg l-1). Motile organisms (fish, portunid crabs, 

stomatopods, penaeid shrimp and squid) are seldom found in bottom waters with oxygen 

concentrations less than 2 mg l-1. Below 1.5 to 1 mg l-1 oxygen concentration, less motile and 

burrowing invertebrates exhibit stress behavior, such as emergence from the sediments, and 

eventually die if the oxygen remains low for an extended period. At minimal concentrations just 

above anoxia, sulfur-oxidizing bacteria form white mats on the sediment surface, and at 0 mg l-1, 

there is no sign of aerobic life, just black anoxic sediments. The composition of the benthic 

communities reflects differences in sedimentary regime, seasonal input of organic material and 

seasonally severe hypoxia/anoxia. Decreases in species richness, abundance and biomass of 

organisms are dramatic when bottom-waters are affected by severe hypoxia/anoxia. Some 

macroinfauna, the polychaetes Ampharete and Magelona and a sipuculan Aspidosiphon, are 

capable of surviving extremely low dissolved oxygen concentrations and/or high hydrogen sulfide 

concentrations. Macroinfauna, primarily opportunistic polychaetes, increase in the spring 

following flux of primary produced carbon, and increase to a lesser extent in the fall following the 

dissipation of hypoxia. Fewer taxonomic groups characterize the severely affected benthos, and 

long-lived, higher biomass and direct-developing species are mostly excluded. Suitable feeding 

habitats (in terms of severely reduced populations of macroinfauna that may characterize 

substantial areas of the seabed) are frequently removed from the foraging base of demersal 

organisms, including the commercially important penaeid shrimps. 

Switzer, Theodore S., Edward J. Chesney, and Donald M. Baltz, 2006. “Habitat Selection 

by Flatfishes along Gradients of Environmental Variability: Implications for 

Susceptibility to Hypoxia in the Northern Gulf of Mexico” Presentation at “Hypoxia 

Effects on Living Resource in the Gulf of Mexico” NOAA Center for Sponsored Coastal 

Ocean Research, Tulane University, New Orleans, LA. September 25 – 26, 2006. 

Abstract: Although eutrophication in the northern Gulf of Mexico contributes to the high fisheries 

productivity characteristic of the region, nutrient over-enrichment leads to the seasonal formation 

of hypoxic (< 2 mg L-1 O2) bottom water along the Louisiana-Texas continental shelf. Despite an 

increase in the magnitude and duration of hypoxic episodes in recent decades, fisheries landings 
Appendix A 

A-6

have remained high; nevertheless, hypoxia remains a persistent threat to the long-term 

sustainability of regional fisheries production. The greatest threat to mobile nekton is likely the 

influence of reduced dissolved oxygen concentrations on habitat quality, potentially forcing the 

movement of individuals and/or prey from generally favorable habitats. At the population level, 

these movements may result in altered spatial distributions that reflect selection of resources along 

gradients of environmental variability. To unravel the potential influence of hypoxia on the 

distribution of nekton, we examined patterns of habitat use by several abundant flatfishes based on 

data collected during summer SEAMAP groundfish surveys from 1987 to 2000. Results from 

habitat suitability analyses indicated that most flatfishes selected a restricted range of suitable 

depths, temperatures, and salinities. Although most flatfishes were tolerant of moderately-low 

dissolved oxygen concentrations, hypoxic environments were generally avoided, indicating that 

hypoxia likely renders large areas of the Gulf of Mexico unsuitable. In comparisons of spatial 

habitat suitabilities between years of moderate (< 15,000 km2) and severe hypoxia (>15,000 km2), 

all flatfishes exhibited a reduction in the suitability of areas immediately west of the Mississippi 

River and a concomitant increase in suitability within adjacent areas. Altered spatial distributions 

corresponded to species-specific suitabilities along depth, temperature, and salinity gradients, 

indicating that habitat suitability analyses may be effective in predicting population-level 

responses to hypoxic episodes. 

Wells, Melissa C., Zhenlin Ju, Sheila J. Heater and Ronald B. Walter, 2006. “Microarray 

Gene Expression Analyses in Medaka (Oryzias latipes) Exposed to Hypoxia” 

Presentation at “Hypoxia Effects on Living Resource in the Gulf of Mexico” NOAA 

Center for Sponsored Coastal Ocean Research, Tulane University, New Orleans, LA. 

September 25 – 26, 2006. 

Abstract: We are investigating the genomic and proteomic effects of hypoxia exposure using the 

Japanese medaka (Oryzias latipes) aquaria fish model as a tool for biomarker discovery. We have 

developed a hypoxia exposure system allowing programmable exposure scenarios and have 

initiated experimental assessment of changes in gene expression and protein abundance using 

microarray and 2D-DIGE gel analyses of hypoxia exposed fish. We present the design, 

construction, validation, and subsequent use of a medaka 8,046 (8K) unigene oligonucleotide 

microarray to begin the study of hypoxia exposure. Array performance was validated via self-self 

hybridization. Optimization of sample size needed for robust array data, based upon the number 

features detected and the signal intensity, suggest 2 μg total RNA as a starting template for 

amplification is sufficient. For treatment, adult medaka are exposed to a hypoxic environment of 

4% dissolved oxygen (DO) for 2 days and then the DO lowered to 2% for an additional 5 days. 

Upon sacrifice, changes in gene expression in brain, liver, skin, and gill tissues of these fish were 

assessed in conjunction with matched control fish exposed similarly to 18% DO. Analyses of 

array results identified 501 features from brain, 442 from gill, and 715 features from liver that 

exhibit statistically significant changes in transcript abundance upon hypoxia exposure. Nine 

features were found to exhibit common expression patterns between all three tissues. Data mining 

of the array results suggest hypoxic exposure results in a general slowdown of metabolic function. 

Real-time PCR was then employed to support the microarray results and this independent 

validation agreed well with the microarray findings. Overall these results indicate the medaka 

microarray will be a sound diagnostic tool for changes in gene expression due to hypoxia 

exposure. 

Zimmerman, Roger J. and James M. Nance, 2001. “Effects of Hypoxia on the Shrimp 

Industry of Louisiana and Texas” Chapter 15 in Rabalais, N.N. and R.E. Turner, Coastal 

Hypoxia: Consequences for Living Resources Coastal and Estuarine Studies, 58 pp 293-

310. 
Appendix A 

A-7

Abstract: This study carries out a statistical test for effects of hypoxia on commercial catch of 

shrimp in the Gulf of Mexico for 1985-97. The analysis combines landings data and interview 

data on fishing effort, catch and location of each trip. The analysis is spatially explicit, based on 

catch in 9 statistical subareas in Louisiana and Texas, with each subarea divided into 10 depth 

zones. Zimmerman and Nance found no correlation of hypoxic area with landings of white shrimp 

or with landings of brown shrimp in Louisiana, but they found a statistically significant 

relationship between hypoxia and combined landings in Texas and Louisiana. The finding of no 

relationship for white shrimp is consistent with prior expectations, because white shrimp are less 

sensitive to hypoxia (Renaud, 1986), and because white shrimp habitat is mostly in-shore the 

hypoxic region. In comparison, brown shrimp travel from inshore areas to offshore in order to 

spawn. Since brown shrimp migrate through the hypoxic region, they are more likely to be 

effected by hypoxia. The absence of a significant relationship between the size of the hypoxic 

region and catch of brown shrimp in Louisiana may be explained by the fact that much of the 

catch in Louisiana occurs in-shore of the hypoxic region, while catch in Texas occurs offshore. 

Zou, Enmin, 2006. “Impacts of Hypoxia on Physiology and Toxicology of the Brown 

Shrimp Penaeus aztecus” Presentation at “Hypoxia Effects on Living Resource in the 

Gulf of Mexico” NOAA Center for Sponsored Coastal Ocean Research, Tulane 

University, New Orleans, LA. September 25 – 26, 2006. 

Abstract: The brown shrimp, Penaeus aztecus, in the northern Gulf of Mexico is faced with dual 

stresses of environmental hypoxia, which occurs as a result of oxygen depletion from microbial 

decomposition of organic materials from algal blooms, and pollution from polycyclic aromatic 

hydrocarbons (PAHs) from petroleum and gas production on the continental shelf of the northern 

Gulf of Mexico. This study aimed to address the questions of 1) whether the presence of PAH 

contamination makes penaeid shrimps more susceptible to hypoxia and 2) whether hypoxia can 

promote PAH bioaccumulation in penaeid shrimps. The susceptibility of shrimps to hypoxia was 

represented by the oxyregulating capacity, a physiological parameter that describes how well an 

animal regulates its oxygen consumption when subjected to hypoxia. It was found that acute 

exposure to naphthalene significantly reduced the oxyregulating capacity of Penaeus aztecus. An 

ensuing consequence of a decrease in oxyregulating ability is that the stress from the lack of 

oxygen would set in sooner in the presence of PAH contamination than when shrimps are in the 

clean environment. Hypoxia was found to have no significant effect on naphthalene 

bioaccumulation in Penaeus aztecus. The absence of a significant effect was attributed to increased 

naphthalene metabolism in the brown shrimp subjected to hypoxia. 
Appendix B 

B-8

B. Appendix B: Flow Diagrams and Mass Balance of Nutrients 

Global Material Cycles 

 For the reader’s information, the following flow diagrams of the global nitrogen, 

phosphorus and silicon cycles are taken from the Encyclopedia of Earth, a new electronic 

reference about the Earth, its natural environments, and their interaction with society. The 

Encyclopedia is a free, fully searchable collection of articles written by scholars, 

professionals, educators, and experts (http://www.eoearth.org/eoe/about). Its contents 

may be freely copied and distributed with proper attribution. These diagrams are not the 

deliberative products of the SAB Panel but are provided to illustrate important processes 

discussed in this report including: fertilization, nitrogen fixation, nitrification, 

denitrification, ammonification, nutrient assimilation, sedimentation, recycling from 

sediment, and weathering of rocks. Chemical equations representing processes depicted 

in the flow diagrams are available from many sources in the published literature 

including standard textbooks on biogeochemistry, limnology, and oceanography.
Appendix B 

B-9

Figure 54: Nitrogen Cycle Flow Diagram. Taken from Encyclopedia of Earth (2007) at 

http:www.eoearth.org/global_material_cycles.
Appendix B 

B-10

Figure 55: Phosphorus Cycle Flow Diagram. Taken from Encyclopedia of Earth (2007) at 

http:www.eoearth.org/global_material_cycles. 
Appendix B 

B-11

Figure 56: Silicon Cycle Flow Diagram. Taken from Encyclopedia of Earth (2007) at 

http:www.eoearth.org/global_material_cycles. 
Appendix B 

B-12

Atmospheric deposition

 The Integrated Assessment concluded that atmospheric deposition as a new 

nitrogen input to the Mississippi River basin was not as important as agricultural sources 

but that deposition nonetheless was a significant source (Goolsby et al., 1999). 

Atmospheric deposition of nitrogen generally shows a trend of increasing from west to 

east in the Mississippi basin, and deposition was a particularly important source of 

nitrogen in the Ohio River basin (Goolsby et al., 1999). The Integrated Assessment

followed the net anthropogenic nitrogen input (NANI) budgeting approach established by 

the International SCOPE Nitrogen Project in assuming that deposition of oxidized 

nitrogen (NOy) is a new input of nitrogen while the deposition of ammonium is not but 

rather is a recycling of nitrogen emitted to the atmosphere from agricultural sources 

within the basin (Howarth et al., 1996). The oxidized nitrogen is presumed to come 

largely from fossil-fuel combustion and, thus, is not accounted for in any other input to 

the budget (Howarth et al., 1996; Goolsby et al., 1999). The Integrated Assessment

further considered that the deposition of organic nitrogen was a new input of nitrogen 

(Goolsby et al., 1999). 

 The Integrated Assessment used monitoring data to estimate NOy deposition and 

made a very rough guestimate for the magnitude of deposition of organic nitrogen. They 

used data from the NADP for wet deposition and from CASTnet for dry deposition. This 

yielded an average estimate of NOy deposition for the Mississippi River basin for the 

time period 1988 to 1994 of 3.4 kg N/ha/yr (3 lb N/ac/yr), of which 2 kg N/ha/yr (1.8 lb 

N/ac/yr) was nitrate in wet deposition and 1.4 kg N/ha/yr (1.25 lb N/ac/yr) was NOy dry 

deposition (Goolsby et al., 1999). The assessment estimated the deposition of organic 

nitrogen as 1 kg N/ha/yr (0.89 lb N/ac/yr), yielding a total estimate for new nitrogen 

deposition of 4.4 kg N/ha/yr (3.9 lb N/ac/yr) (Goolsby et al., 1999). This can be 

compared with an estimate for NOy deposition derived from the GCTM model, which 

estimates deposition rates from data on emissions to the atmosphere and on rates of 

reaction and advection within the atmosphere (Prospero et al., 1996). For the Mississippi 

River basin for essentially the same time period used in the Integrated Assessment, the 

GCTM model suggested a total NOy deposition of 6.6 kg N/ha/yr (5.9 lb N/ha/yr), with 

6.2 kg N/ha/yr (5.5 lb N/ac/yr) of this input being attributable to new inputs from fossilfuel burning and 0.4 kg N/ha/yr (0.36 lb N/ac/yr) originating from natural sources 

(Howarth et al., 1996). 

 Holland et al. (1999, 2005) noted that deposition estimates based on monitoring 

data are typically lower than those from emission-based models across most of the United 

States. For the northeastern United States from Maine through Virginia, the estimates 

from the GCTM model (Howarth et al., 1996) are again almost twice as high as are 

estimates from NADP and CASTnet monitoring data (Boyer et al., 2002). There are 

many possible reasons for this discrepancy, but probably at least part of the problem lies 

with an underestimation of dry deposition by the CASTnet program (Holland et al., 1999; 

Howarth et al., 2006b; Howarth, 2006). Most CASTnet monitoring stations are 

purposefully located away from emission sources, and deposition is likely to be higher 

near these emission sources, creating a bias in the network. Further, the CASTnet 
Appendix B 

B-13

program only estimates deposition of nitrogen in particles and deposition of nitric acid 

vapor. The deposition of several other gases (including NO, NO2, and nitrous acid vapor) 

is not measured. Deposition of these gases, which would be included in the estimates 

from the emission-based models, is likely to be particularly high near emission sources 

(Howarth, 2006). Both the GCTM and TM3 models only estimate deposition at coarse 

spatial scales, but a new emission-based model (CMAQ) shows promise for estimation at 

relatively fine spatial scales (Robin Dennis, NOAA, personal communication). Note that 

this model suggests very high NOy deposition rates near urban centers in the eastern US 

and associated with power plant emissions in the Ohio River basin (Figure 57). 
Appendix B 

B-14

In the mass balance presented in Section 3.2, deposition was estimated as in Goolsby et 

al. (1999). Organic N was not included, however, as there it is unclear what the 

importance is of this form of N, or what an appropriate estimate would be (Keene et al., 

2002). A comparison was made of deposition inputs by region of the NOy estimate used 

in the mass balance and deposition from the CMAQ model for 2001. For the upper 

Mississippi basin, NOy deposition was 4.2 kg N/ha/yr (3.8 lb N/ac/yr), the same as the 

CMAQ model1

. For the Missouri basin both methods again gave similar estimates, with 

NOy deposition of 2.2 N/ha/yr (2 lb N/ac/yr), and CMAQ modeled deposition 2.1 kg 

N/ha/yr (1.9 lb N/ac/yr). For regions with more fuel combustion, the pattern was 

different, with an Ohio basin NOy estimate of 5.0 N/ha/yr (4.5 lb N/ac/yr), and the 

CMAQ model estimate of 8.8 kg N/ha/yr (7.8 lb N/ac/yr). For the lower Mississippi 

River basin, NOy was 3.7 kg N/ha/yr (3.3 lb N/ac/yr), and the CMAQ estimate 5.1 kg 

N/ha/yr (4.6 lb N/ac/yr). Overall, this supports mass balance analysis that for the upper 

Mississippi basin, atmospheric deposition is a small component of N inputs (about 8% of 

N inputs) and is more important in the Ohio region (about 16% of N inputs using the 

CMAQ model for 2001). 

1

CMAQ model unpublished results courtesy of Robin Dennis, NOAA, with analysis by states provided 

by Dennis Swaney, Cornell University; unpublished.

Figure 57: Annual average deposition of NOy across the United States (kg N /hectare-year) 

based on beta-testing runs of the CMAQ model. Note the very high rates of deposition in the 

Ohio River basin. Courtesy of Robin Dennis, NOAA.
Appendix C 

C-15 

C. Appendix C: EPA’s Guidance on Nutrient Criteria 

 In 2000, EPA recommended criteria to States and Tribes for use in establishing 

their water quality standards consistent with section 303(c) of the Clean Water Act 

(CWA). Under section 303(c) of the CWA, States and authorized Tribes have the 

primary responsibility for adopting water quality standards as State or Tribal law or 

regulation. The standards must contain scientifically defensible water quality criteria that 

are protective of designated uses. On its website at 

http://www.epa.gov/waterscience/criteria/nutrient/ecoregions/, EPA provides 

recommended criteria for nutrients in four major types of waterbodies – lakes and 

reservoirs, rivers and streams (U.S. EPA, 2006b), estuarine and coastal areas, and 

wetlands – across fourteen major ecoregions of the United States. The SAB Panel asked 

EPA for a comparison of the SAB Panel’s proposed 45% reductions for TN and TP flux 

to the nutrient levels that would correspond to EPA’s recommended ecoregional criteria. 

 Before presenting that preliminary analysis, the following caveats are stressed. 

• EPA’s recommended ecoregional nutrient criteria are not laws or 

regulations; they are guidance that States and Tribes may use as a starting 

point for developing criteria for their water quality standards. Final 

criteria developed by States and Tribes may have concentrations higher or 

lower than EPA ecoregional recommendations, or, if scientifically 

defensible, not include a nutrient if an impact on “designated use” was not 

found. 

• EPA’s recommended ecoregional nutrient criteria do not take into account 

local site-specific conditions and “designated uses” for particular water 

bodies (e.g., recreation, water supply, aquatic life, agriculture). 

• EPA’s guidance for ecoregional nutrient criteria are based on ambient 

concentrations of nutrients (expressed in mg/L or ug/L) in various 

ecoregions. By contrast, the SAB Panel’s recommended reductions of TN 

and TP are based on flux (expressed in million metric tons of TN and TP 

discharged at the mouth of the Mississippi River). A direct comparison of 

concentrations to flux necessitates the simplifying assumption that 

percentage reductions in concentrations have a one-to-one correspondence 

with percentage reductions in flux. 

• EPA’s guidance for ecoregional criteria is based on estimated “reference 

conditions” i.e., reference sites chosen to represent the least culturally 

impacted waters of the class existing at the present time. The estimated 

reference conditions are based on the 25th percentile of the frequency 

distribution of nutrient concentration data available for each ecoregion. 

This assumption lends uncertainty to EPA’s guidance for ecoregional 

nutrient criteria. 
Appendix C 

C-16 

 Given these caveats, the following analysis by EPA Office of Water’s Office of 

Science and Technology and EPA’s Office of Research and Development allows some 

comparison between EPA’s guidance for ecoregional nutrient criteria and the SAB 

Panel’s proposed 45% nutrient reductions. 

Comparison of SAB Nitrogen and Phosphorus Recommendations with EPA Nitrogen 

and Phosphorus Criteria Recommended Reference Conditions – Submitted by EPA’s 

Office of Water, 8-24-07. 

Question: How do the 45% recommended reductions in nitrogen (N) and phosphorus 

(P) at the mouths of the Mississippi and Atchafalaya Rivers compare with the 25th 

percentile of TN and TP concentration data from ecoregions draining the MississippiAtchafalaya River Basin (MARB)? 

Answer: This question is addressed with a preliminary approach. A more thorough 

approach is needed, but this would require a longer period of time. 

 The preliminary approach was developed by staff from the EPA Office of 

Research and Development’s Gulf Breeze Lab and the EPA Office of Water’s Office 

of Science and Technology using USGS loading estimates from the lower Mississippi 

River at St. Francisville, LA and the Atchafalaya River at Melville, LA over the past 

20 years. This approach compares the 45% reduction in nitrogen and phosphorus 

recommended by the SAB, to the 25th percentiles of the distribution of data in EPA’s 

National Nutrient Database for total nitrogen (TN) and total phosphorus (TP) in each 

aggregate nutrient ecoregion of the MARB. These 25th percentiles represent EPA’s 

approximated reference conditions for those ecoregions. 

 It is important to note that these 25th percentile values are not intended to be 

implemented or promulgated directly as criteria. Rather, EPA developed the nutrient 

criteria recommendations with the intent that they serve as a starting point for States 

and Tribes to develop more refined criteria, as appropriate, to reflect local conditions. 

States and Tribes may adopt criteria that are higher or lower than these 25th 

percentiles. Text in two EPA documents help clarify the use of the ecoregional 

reference condition values. See introductions to the ecoregional criteria documents at 

http://www.epa.gov/waterscience/criteria/nutrient/ecoregions/rivers/index.html and 

EPA’s Nutrient Criteria Technical Guidance Manual for Rivers and Streams 

(http://www.epa.gov/waterscience/criteria/nutrient/guidance/rivers/chapter_1.pdf). 

 Given this description, one can compare a 45% reduction in N and P measured 

in two locations to the estimated reference conditions in each of the MARB 

ecoregions to obtain a rough estimate of whether a 45% reduction could be more or 

less stringent than what could result if EPA’s recommended reference conditions 

were adopted without further modification, as state water quality standards. 

Data Sources: River flow and nutrient flux are monitored and computed by the U.S. 

Geological Survey’s (USGS) National Stream Quality Accounting Network 
Appendix C 

C-17 

(NASQAN) program at numerous river gauge stations in the Mississippi River Basin. 

A description of the USGS NASQAN program, the flux estimation methodology and 

the downloadable data records are available at http://toxics.usgs.gov/hypoxia/. 

Monthly average nutrient concentrations were calculated from the USGS data as 

The Monthly Average Nutrient Concentration = USGS Monthly Load/ Monthly 

Average discharge rate where monthly average discharge rates for the mainstem 

Mississippi River were calculated from daily discharge rates obtained at the Tarbert 

Landing, MS gauge (ID = 01100). 

 

Nitrogen. The median monthly nitrate concentration for the combined Mississippi 

River at St. Francisville and Atchafalaya River at Melville over the period 1979 – 

2007 is 1.24 mg/L. In comparison, historical data from the Mississippi River at St. 

Francisville indicate that the median nitrate concentrations during the period 1955-

1970 was 0.6 mg/L. 

Nitrate, as a component of TN is about 60% on average (based on USGS nutrient load 

data); thus 1.24 mg/L nitrate would extrapolate to 2.07 mg/L TN. 

A proposed 45% reduction of 2.07 mg/L TN would yield a concentration of 1.14 

mg/L TN. 

The relevant EPA recommended ecoregional reference conditions for TN are: 

Ecoregion IV - 0.56 mg/L 

Ecoregion V - 0.88 mg/L 

Ecoregion VI – 2.18 mg/l 

Ecoregion VII - 0.54 mg/L 

Ecoregion IX - 0.69 mg/L 

Ecoregion X - 0.76 mg/L 

Ecoregion XI - 0.31 mg/L 

These values range from 27% to 191% of the estimated 1.14 mg/L TN that would 

result from a 45% reduction, with all but one value below 100% (the Corn Belt and 

northern Great Plains ecoregion VI). This suggests that a 45% reduction of estimated 

median monthly TN concentrations to 1.14 mg/L would likely be less stringent than 

could be obtained if states adopted EPA’s recommended reference condition values 

into state water quality standards for TN. 

Phosphorus. Using the same data (Mississippi River at St. Francisville and 

Atchafalaya River at Melville, 1979-2007, monthly means), the median monthly 

concentration of TP is 202 ug/L. Thus a 45% reduction of 202 ug/L TP would yield a 

concentration of 111 ug/L. 

The relevant EPA recommended ecoregional reference conditions for TP are: 
Appendix C 

C-18 

Ecoregion IV – 23.00 ug/L 

Ecoregion V – 67.00 ug/L 

Ecoregion VI – 76.00 ug/L 

Ecoregion VII – 33.00 ug/L 

Ecoregion IX – 36.00 ug/L 

Ecoregion X – 128.00 ug/L 

Ecoregion XI – 10.00 ug/L 

These values range from 9% to 115% of the estimated 111 ug/L TP that would result 

from a 45% reduction, with all but one value below 100% (the Texas-Louisiana 

Coastal and Mississippi Alluvial Plains ecoregion X). This also suggests that a 45% 

reduction of estimated median monthly TP concentrations to111 ug/L would likely be 

less stringent than could be obtained if states adopted EPA’s recommended reference 

condition values into state water quality standards for TP. 

A More Comprehensive Approach 

 A thorough comparison of the distribution approach to reference condition 

estimation and the 45% reduction in TN and TP could be made by calculating the 

nutrient concentrations from the USGS loading estimates at river gauge stations at 

each of the nine subbasins. The USGS provides monthly or annual nutrient flux 

estimates and river flow data from which nutrient concentration data can be derived 

(http://toxics.usgs.gov/hypoxia/.) These data provide values over many years for 9 

subbasins located within the MARB. The data could be used in the following steps to 

compare the two sets of values: 

• Use the USGS nutrient loading data to compile a TN and TP concentration 

dataset for each subbasin; 

• Calculate the median TN and TP concentrations at each of the nine subbasin 

river gauge stations; 

• Overlay nutrient ecoregions on subbasins and extract nutrient ecoregional data 

from subbasins. From this refined data set, calculate the median value of the 

seasonal 25th percentiles of TN and TP for the ecoregion-subbasin. 

These data can be used for the following comparisons: 

1) Calculate the concentrations resulting from a 45% reduction in the median 

concentration for each subbasin. 

2) Compare these to the EPA 25th percentiles (ecoregional reference conditions) in 

each subbasin, or specific subbasins of interest. 

Submitted by EPA’s Office of Water, 8-24-07. 
Appendix D 

D-19 

D. Appendix D: Calculation of Point Source Inputs of N and P 

 As discussed in Sections 3.2 and 4.5.8, estimates of N and P fluxes from sewage 

treatment plants from the MART (2006b) report were much lower for both total N and 

total P than in the Integrated Assessment. As pointed out in the MART report, much of 

this decline is thought to be due to the values assigned for total N and P concentrations in 

sewage treatment plants effluent. Few measured data were used, but rather estimated 

values were applied to most. Most estimates were made using a “typical pollutant 

concentration” (TPC) for N or P based on the level of treatment. These TPC’s were from 

an update of values compiled in a report by Tetra Tech (1998). The 2006 MART report 

assumed that sewage treatment plants with advanced wastewater treatment had TPC 

values of 5.6 and 0.82 mg/L for total N and total P, respectively. The MART report then 

applied these assumed values to estimated daily discharges to calculate an estimated daily 

flux. The MART report further assumed that plants that had less than advanced 

wastewater treatment had TPC values of 11.2 and 2.02 mg/L for total N and total P, 

respectively, applied to estimated daily discharges to calculate an estimated daily flux. 

The Panel is not comfortable with these assumptions and instead believes that most 

wastewater treatment plants in the MARB had TPC’s applied that were too low. The 

Panel, therefore, adjusted the database, by using TPC’s of 11.2 and 2.02 mg/L for total N 

and total P, respectively, for plants with advanced wastewater treatment, and TPC’s of 15 

and 4 mg/L for total N and total P, respectively, for plants with less than advanced 

treatment. 

 As an example of how these adjustments changed estimates, the Panel examined 

seven Chicago plants (Stickney is the largest sewage treatment plant in the basin) and one 

in Champaign-Urbana, IL, where measured flux data were available (daily to weekly 

measurements of total N and total P and flow were made at each plant). From this 

analysis, it is clear that the TPCs used in the MART report were not appropriate and gave 

substantially lower flux estimates (Table 20) than the actual measured values. The 

MART report indicated that each of these plants had advanced treatment, and therefore 

applied their estimated TPCs of 5.6 mg/L total N and 0.82 mg/L total P respectively. 

Most plants in the MARB do not have treatment processes (either biological or chemical) 

to remove P, and much of the advanced treatment is to nitrify ammonium to nitrate, 

because most are permitted for ammonia in effluent. 
Appendix D 

D-20 

Table 20: Comparison of MART estimated sewage treatment plant annual effluent loads of total N and P 

and values from measurements at each plant for 2004. 

MART Measured MART Measure

d 

Plant 

tons N/yr 

% Diff 

tons P/yr 

% Diff 

Stickney 6,282 9,850 64 921 1,105 83 

Calumet 1,799 (3,599) 3,243 55 264 (650) 1,065 25 

Lemont 13 (26) 51 25 2 (5) 8 23 

Northside 2,259 (4,518) 3,161 71 331 (207) 441 75 

Egan 209 (418) 386 54 31 (75) 99 31 

Hanover Park 70 (139) 144 48 10 (73) 33 31 

Kirie 201 (402) 331 61 29 (73) 44 67 

Champaign-Urbana 77 (155) 310 25 11 (28) 58 20 

* All plants are in Illinois. Also shown in red is the recalculated MART value as described below, except 

for Stickney, where actual values were used because of plant size and concentration considerations. 

 This analysis supports the Panel’s use of increased TPC’s when estimating point 

source loads. Therefore, all plants that were labeled as advanced treatment and used the 

Clean Water Needs Survey data for load estimates were recalculated using total N and P 

concentrations of 11.2 and 2.02 mg/L, respectively (this included most plants, and some 

were estimated using the permit compliance system data and were not recalculated). 

These concentrations were much closer to the values reported by the plants in Table 20 

(red values in table), although there still was considerable variability, and included 2,080 

point sources (the total database has 33,302 point sources of all types). For plants 

identified as receiving secondary treatment, total N and P concentrations of 15 and 4 

mg/L, respectively, were applied (there were 4,480 plants of this type). For the seven 

plants in Table 20 recalculated this way, total N and P fluxes were 113 and 91% of 

measured values, respectively, much closer to the measured values than the original 

MART values. The Panel’s discussion of point sources in the MARB utilizes these 

adjustments to the MART values. Finally, the Panel again emphasizes that measured 

data are generally not available in these large databases, so that many assumptions need 

to be made.
Appendix E 

E-21 

E. Appendix E: Animal Production Systems 

Intensification of animal feeding operations 

 Current census information shows that there has been an 18% increase in the 

number of pigs in the U.S. over the last 10 years along with a 72% decrease in the 

number of farms. Over the same 10 years, the number of dairies has decreased by 40%, 

but herd size has increased by 50%. A similar trend in the poultry and beef industries has 

also occurred, with 97% of poultry production in the U.S. coming from operations with 

more than 100,000 birds and over a third of beef production from <2% of the feedlots 

(Gardner, 1998). Fattened cattle numbers remained fairly constant from 1982 to 1997 but 

the number of fattening operations decreased over 50 percent (Kellogg et al., 2000). 

Overall, cattle, pig, and poultry numbers have increased 10% to 30%, while the number 

of farms on which they were reared has decreased 40% to 70% over the last 10 years 

(Gardner 1998). 

Nutrient budgets 

 The large-scale consolidation has created much larger animal feeding operations, 

which makes economical utilization and re-distribution of manure to croplands difficult 

and has profound consequences for farm and regional nutrient transfer and management 

within the MARB. For example, the accumulation of nutrients is first evident at the farm 

scale, where N and P management is affected by daily operation decisions and the longterm goals of each farmer. For example, the potential for P and N surplus on farms with 

AFOs can be much greater than in cropping systems where nutrient inputs become 

dominated by feed rather than fertilizer (Table 21). With a greater reliance on imported 

feeds, only 30% of N and 29% of P in purchased feed for a 1280-hog operation on a 30-

ha farm could be accounted for in farm outputs. These nutrient budgets clearly show that 

animal feed is the largest input of nutrients to farms with AFOs, and thus is the primary 

source of on-farm nutrient excess, for which a resolution will require innovative 

management. Current animal number and estimated manure N and P production within 

the MARB is given in Table 22. 
Appendix E 

E-22 

Table 21: Farming System and Nutrient Budget. 

Nutrient input in 

Farming system 

Feed Fertilizer 

Output in 

produce Surplus Nutrient 

utilization 

 - - - - - - - - - - - - - - kg ha-1 yr-1 - - - - - - - - - - - - - - - - % 

Nitrogen budget 

Cash crop a

 -- 95 92 3 97 

Dairy b

 155 40 75 120 38 

Hog c

 390 10 120 280 30 

Poultry d

 5800 -- 1990 3810 34 

Phosphorus budget 

Cash crop a

 -- 22 20 2 91 

Dairy b

 30 11 15 26 37 

Hog c

 105 -- 30 75 29 

Poultry d

 1560 -- 440 1120 28 

a

 30 hectare cash crop farm growing corn and alfalfa. 

b

 40 hectare farm with 65 dairy Holsteins averaging 6600 kg milk cow-1 yr-1, 5 dry cows and 35 heifers. 

Crops were corn for silage and grain and alfalfa and rye for forage. c

 30 hectare farm with 1280 hogs; surplus includes 36 kg P and 140 kg N ha-1 yr-1 manure exported from 

the farm. 

d

 12 hectare farm with 74,000 poultry layers; surplus includes 180 kg P and 720 kg N ha-1 yr-1 manure 

exported from the farm. 
Appendix E 

E-23 

Table 22: Number of animals and amount of manure produced and N and P excreted within the MARB 

states based on information from the 1997 U.S. Census of Agriculture (data obtained from USDA-ERS, 

http://ers.usda.gov/data/MANURE/). 

Animal 

type # Farms Animal Units Manure 

excreted † 

Manure N 

excreted 

Manure P 

excreted 

 Mg - - - - - million kg - - - - - 

Beef 837,972 52,627,536 71,354,009 2,712 864 

Dairy 77,363 5,944,742 13,287,687 439 79 

Poultry 45,870 3,044,000 8,743,736 433 149 

Swine 84,717 6,591,998 7,310,054 419 124 

† Manure in dry state, as excreted adjusted for water content. 

Nutrient Surpluses 

 USDA (2003) estimated the amount of manure produced from animal distribution 

numbers from the 1997 U.S. Census of Agriculture, using standard values of manure 

production and nutrient concentration for each animal type. Estimates of excess N and P 

were calculated based on crop N and P removal and the assumption that all suitable crop 

and pasture land was available for manure application (Figure 43 and Figure 44). Most 

areas with CAFOs have some excess N (Figure 43) and P (Figure 44). These 

distributions demonstrate that within the MARB, regional excesses were similar for N 

and P. 

Targeting Remedial Strategies Within the MARB 

 The importance of targeting nutrient management within a watershed is shown by 

several MARB studies. In the early 1980’s conservation practices were installed on 

about 50% of the Little Washita River watershed (54,000 ha or 133,000 ac) in central 

Oklahoma. Practices included construction of flood control impoundments, eroding gully 

treatment, and conservation tillage (Sharpley and Smith, 1994; Sharpley et al., 1996). 

Although conservation measures decreased N and P export 5 to 13 fold, there was no 

effect on P concentration in flow at the outlet of the main Little Washita River watershed. 

Thus, a lack of effective targeting of nutrient management and control of major sources 

of nutrient export contributed to field or subwatershed scale responses not being 

translated to reductions in nutrient export from the main Little Washita River watershed. 
Appendix E 

E-24 

Managing Manures 

 Manure application timing and method relative to rainfall influences the 

concentration of N and P in runoff (Dampney et al., 2000; Sims and Kleinman, 2005). 

For example, several studies have shown a decrease in N and P loss with an increase in 

the length of time between manure application and surface runoff (Djodjic et al., 2000; 

Edwards and Daniel, 1993a; Sharpley, 1997; Westerman et al., 1983). This decrease can 

be attributed to the reaction of added P with soil and dilution of applied P by infiltrating 

water from rainfall that did not cause surface runoff. 

 The incorporation of manure into the soil profile either by tillage or subsurface 

placement decreases the potential for P loss in surface runoff. Rapid incorporation of 

manure also reduces NH3 volatilization and potential loss in runoff as well as improving 

the N:P ratio for crop growth. Mueller et al. (1984) showed that incorporation of dairy 

manure by chisel plowing reduced total P loss in runoff from corn 20-fold, compared to 

no-till areas receiving surface applications. In fact, P loss in runoff was decreased by a 

lower concentration of P at the soil surface and a reduction in runoff with incorporation 

of manure (Mueller et al., 1984; Pote et al., 1996). As with fertilizer application 

methods, other factors are important in selecting or recommending the most appropriate 

application method. Equipment availability, whether the soil is sufficiently free of rocks 

to allow subsurface application, labor requirements, product availability, and availability 

of operating capital all affect the application method decision. 

Crop selected to receive manure application 

 Manure has traditionally been applied for corn or other grass production. 

However, corn acreage to which manure is applied has not expanded proportionally to 

animal operation expansions; thus the risks increase for applying manure in excess of the 

amount necessary to meet crop nutrient requirements (Schmitt et al., 1996; Dou et al., 

1998). One solution to minimize these risks, and the subsequent potential risk of NO3

leaching to ground water, is to select alternative crops to receive manure applications. 

Although legumes are not usually considered for manure application, soybean can 

annually remove as much as 385 kg N/ha (344 lb N/ac) (Shibles, 1998) and alfalfa as 

much as 500 kg N/ha (446 lb N/ac) (Russelle et al., 2001), compared to less than 200 kg 

N/ha (179 lb N/ac) for corn. Schmidt et al. (2000) demonstrated that nodulation in 

soybean effectively compensated with additional N when manure N was insufficient to 

meet crop demands; so if necessary, manure could be applied conservatively without risk 

of applying too little to meet crop needs. 

Rate and frequency of application 

 As might be expected, N and P loss in runoff increases with greater frequency and 

rates of applied manure (Edwards and Daniel, 1993b; McDowell and McGregor, 1984). 

Although rainfall intensity and duration, as well as when rainfall occurs relative to 

applied manure, influence the concentration and overall loss of manure N and P in runoff, 

the relationship between potential loss and application rate is critical to establishing 
Appendix E 

E-25 

environmentally sound nutrient management guidelines. Also evident is that the effect of 

applied manure on increasing the concentration of P in surface runoff can be long lasting. 

For instance, Pierson et al., 2001 found that a poultry litter application tailored to meet 

pasture N demands elevated surface runoff P for up to 19 months after application. 

Although few studies have evaluated the loss of P in surface runoff as a function of 

application frequency, more frequent manure applications can be expected to rapidly 

increases soil P (Haygarth et al., 1998; Sharpley et al., 1993; 2005; Sims et al., 1998), 

with a concomitant increases in runoff P loss. 

Intensity and duration of grazing 

 As beef grazing of pastures is an important component of animal production in 

many regions of the MARB, careful management of grazing is needed to minimize P loss 

and water quality impacts. The localized accumulations of P where manure is deposited 

can saturate the P sorption capacity of a soil, increasing the potential for P loss from 

grazed pastures in runoff or drainage waters. However, at a field and watershed scale, it 

is likely that critical stocking factors, such as density and duration, will influence both 

hydrologic and chemical factors controlling P transport. For example, Owens et al. 

(1997) found that decreasing grazing density and duration dramatically reduced runoff 

and erosion from a pastured watershed in Ohio. Clearly, increased runoff and erosion 

with grazing will enhance the potential for P loss. In Oklahoma, Olness et al. (1975) 

found that P losses were greater from continuously (4.6 kg P/ha/yr or 4.1 lb P/ha/yr) than 

rotationally grazed pastures (1.3 kg P/ha/yr or 1.2 lb P/ha/yr). In fact, P losses with 

continuous grazing were greater than from alfalfa or wheat (2.7 kg P/ha/yr (2.4 lb 

P/ha/yr); Olness et al., 1975). However, the work of Owens et al. (1997) does show that 

when management is changed, the impacts of the previous grazing impacts were not long 

lasting, changing within a year. Even so, there is a need to determine critical stocking 

densities and durations as a function of grazing management. 

Stream-bank fencing 

 By observing four pastured dairy herds with stream access over four intervals 

during the spring and summer of 2003 in the Cannonsville Watershed south central, New 

York, James et al. (2007) were able to estimate fecal P contributions to streams. In the 

herds observed, on a per cow basis, cattle were especially likely to defecate in the stream, 

although they spent a small proportion of their time there. On average, approximately 

30% of all fecal deposits expected from a herd were observed to fall on land within 40-m 

of a stream, and 7% fell directly into streams. Although amenities in pasture (such as 

water troughs, feeders, salt, and shade located away from the stream) did affect where 

cattle congregated, the stream demonstrated a consistent draw. 

 Using spatial databases of streams, pasture boundaries, and animal characteristics 

(i.e., number of cattle, time in pasture, and type of cattle [heifers versus milk cows]) for 

90% of the dairy farms in the Cannonsville watershed, approximately 3,600 kg (7,940 lb) 

of manure P are estimated as deposited directly into streams with 7,650 kg (16,900 lb) 

deposited in pasture near streams (<10 m) from the 11,000 dairy cattle in the watershed. 
Appendix E 

E-26 

At this magnitude, P loadings represent a significant environmental concern, with instream deposits equivalent to approximately 12% of watershed-level P loadings attributed 

to agriculture (Scott et al., 1998). Riparian shade can also attract grazing cattle and 

influence P loss in stream flow.