Science Advisory Board 

Office of the Administrator

SABReactive Nitrogen in the United States: 

An Analysis of Inputs, Flows, 

Consequences, and Management Options

EPA-SAB-11-013 | August 2011 | www.epa.gov/sab

A Report of the EPA Science Advisory Board 
Cover photos by Eric Vance and other photographers, U.S. EPA, courtesy of U.S. EPA.
i

The Honorable Lisa P. Jackson 

Administrator 

U.S. Environmental Protection Agency 

1200 Pennsylvania Avenue, N.W. 

Washington, D.C. 20460

Subject: Reactive Nitrogen in the United States: An Analysis of Inputs, Flows, Consequences, and Management 

Options – A Report of the EPA Science Advisory Board 

Dear Administrator Jackson: 

Excess reactive nitrogen compounds in the environment are associated with many large-scale environmental concerns, 

including eutrophication of surface waters, toxic algae blooms, hypoxia, acid rain, nitrogen saturation in forests, and 

global warming. In addition, reactive nitrogen is associated with harmful human health effects caused by air pollution 

and drinking water contamination. Reactive nitrogen (hereafter referred to as Nr) includes all biologically active, 

chemically reactive, and radiatively active nitrogen compounds in the atmosphere and biosphere of the earth, in contrast 

to non-reactive gaseous N2. EPA and other federal and state agencies have implemented programs to reduce the risks 

posed by excessive Nr, but a more comprehensive and integrated approach is needed to manage the use of Nr in a way 

to achieve its benefits, such as fertilizer for food production, and mitigate its damages as it is introduced to and cycles 

repeatedly through the environment in different forms and media. 

The Science Advisory Board (SAB) Integrated Nitrogen Committee has conducted a study to analyze sources and 

fate of Nr in the United States and provide advice to EPA on integrated nitrogen research and control strategies. We are 

pleased to submit the SAB report, Reactive Nitrogen in the United States: An Analysis of Inputs, Flows, Consequences, 

and Management Options. Our objectives for this study were to:

• Identify and analyze from a scientific perspective the problems Nr presents in the environment and the links among 

them;

• Evaluate the contribution an integrated nitrogen management strategy could make to environmental protection;

• Identify additional risk management options for EPA’s consideration; and

• Make recommendations to EPA concerning improvements in nitrogen research to support risk reduction.

The SAB report provides findings and recommendations addressing these study objectives. Assessment of the challenges 

and costs to the Agency of implementing the recommendations is beyond the scope of the report. 

In general, the SAB finds that:

• In the United States, human activities across multiple sources currently introduce more than five times the Nr into 

the environment than natural processes. The largest U.S. sources of new Nr entering the U.S. environment include: 

the creation and use of synthetic fertilizers, Nr created by legumes, and the combustion of fossil fuels.

• Much of the Nr used to ensure a plentiful supply of food, fiber and biofuel is released to the environment, as is the 

Nr formed during fossil fuel combustion.

• The introduction of human created Nr into the environment degrades air and water quality, which can cause harmful 

algae blooms, hypoxia, fish kills, loss of drinking water potability, loss of biodiversity, forest declines, and human 

health problems resulting in losses of billions of dollars per year. 

UNITED STATES ENVIRONMENTAL PROTECTION AGENCY

WASHINGTON, DC 20460

August 18, 2011

OFFICE OF THE ADMINISTRATOR

SCIENCE ADVISORY BOARD

EPA-SAB-11-013
ii

• Multiple strategies and actions exist to more effectively minimize the inputs of Nr to the environment and maximize 

nitrogen use efficiency. 

The SAB provides the following overarching recommendations to improve the management of Nr. 

• The framing of the movement of nitrogen among various environmental reservoirs in terms of the nitrogen 

cascade concept provides a means for tracking nitrogen as it changes form and passes through multiple ecosystems 

and media. Given this complexity, innovative management systems and regulatory structures reflecting these 

characteristics of Nr are required to address the significant environmental and human health damage caused by Nr. 

New institutional structures and relationships that also reflect the multi-media and multi-form character of Nr and its 

flows and transformations through the environment will have to be created for effective control and management.

• The SAB recommends an integrated approach to the management of Nr. This approach must use a combination 

of implementation mechanisms appropriate to the specific environmental and policy contexts and supported by 

critical research on the specific risks of Nr and on decreasing the risks of Nr. The approach must reflect an integrated 

policy that recognizes the complexity and trade-offs associated with the nitrogen cascade while recognizing that 

intervention points vary in terms of efficiency and cost effectiveness.

• EPA should form an intra-Agency Nr management task force that will build on the existing breadth of Nr research 

and management capabilities within the Agency. Its objective should be to increase scientific understanding of: 

(1) Nr impacts on terrestrial and aquatic ecosystems, human health, and climate; (2) Nr-relevant monitoring 

requirements; and (3) the most efficient and cost effective means by which to decrease various adverse impacts of Nr 

loads as they cascade through the environment.

• Successful Nr management will require changes in the way EPA interacts with other agencies. The SAB 

recommends that EPA convene a reactive nitrogen inter-agency management task force with broad representation 

from other agencies and departments involved with Nr control or utilization. This is essential to coordinate 

federal programs that address Nr concerns and would help ensure clear responsibilities for monitoring, modeling, 

researching, and managing Nr in the environment. Similar efforts at coordination and joint action need to be made 

among and between agencies at both the state and federal level.

In the context of addressing the specific study objectives, the SAB explored how an estimated 25 percent reduction 

in Nr introduced into the environment might be achieved with existing technology in the coming 10 to 20 years through 

actions that could be taken by EPA, other management authorities, and other public and private organizations. Specific 

actions include increased controls of oxides of nitrogen, improved reactive nitrogen uptake by agricultural crops, largescale creation and restoration of wetlands for nitrogen removal in agricultural landscapes with high Nr in surface waters, 

decreased loss of reactive nitrogen from agricultural lands and animal feeding operations, and decreased discharge of 

reactive nitrogen from point sources and developed (urban) lands. However, dealing effectively with reactive nitrogen’s 

cascade through air, water and land will require an integrated management approach that is multi-media and multistressor as suggested by recent initiatives by EPA’s Office of Research and Development.

The most important task for EPA and allied agencies and departments will be to effectively inform the public of the 

costs and dangers of excess Nr. Without strong public support, the widespread efforts necessary to control Nr will not be 

possible.

In closing, we appreciate the opportunity to provide advice on this very important topic, and we look forward to 

receiving your response. The SAB stands ready to provide more information as it may be useful and would be pleased to 

assist EPA in the implementation of the report’s recommendations, if the EPA would find such support valuable.

 Sincerely,

 Dr. Otto C. Doering III

 Chair

 Integrated Nitrogen 

 Committee

 Science Advisory Board

Dr. James N. Galloway

Chair (2007-2009)

Integrated Nitrogen Committee

Science Advisory Board

Dr. Thomas L. Theis

Vice-Chair (2008-2009)

Integrated Nitrogen Committee

Science Advisory Board

Dr. Deborah L. Swackhamer 

Chair

Science Advisory Board
iii

CHAIR

Dr. Otto C. Doering III, Professor, Department of Agricultural Economics, Purdue University, W. Lafayette, IN

CHAIR (2007-2009)

Dr. James N. Galloway, Sidman P. Poole Professor of Environmental Sciences, Associate Dean for the Sciences, 

College and Graduate School of Arts and Sciences, University of Virginia, Charlottesville, VA

VICE-CHAIR (2008-2009)

Dr. Thomas L. Theis, Director, Institute for Environmental Science and Policy, University of Illinois at Chicago, 

Chicago, IL

MEMBERS

Dr. Viney Aneja, Professor, Department of Marine, Earth, and Atmospheric Sciences, School of Physical and 

Mathematical Sciences, North Carolina State University, Raleigh, NC

Dr. Elizabeth Boyer, Associate Professor, School of Forest Resources and Assistant Director, Pennsylvania State 

Institutes of Energy & the Environment, and Director, Pennsylvania Water Resources Research Center, Pennsylvania 

State University, University Park, PA

Dr. Kenneth G. Cassman, Professor, Department of Agronomy and Horticulture, Institute of Agriculture and Natural 

Resources, University of Nebraska, Lincoln, NE

Dr. Ellis B. Cowling, University Distinguished Professor At-Large Emeritus, Colleges of Natural Resources and 

Agriculture and Life Sciences, North Carolina State University, Raleigh, NC

Dr. Russell R. Dickerson, Professor, Department of Atmospheric and Oceanic Science, The University of Maryland, 

College Park, MD

Mr. William Herz, Vice President for Scientific Programs, The Fertilizer Institute, Washington, DC

Dr. Donald L. Hey, Executive Director, Wetlands Research, Inc, Chicago, IL

Dr. Richard Kohn, Professor, Animal Sciences Department, University of Maryland, College Park, MD

Dr. JoAnn S. Lighty, Chair and Professor, Chemical Engineering, University of Utah, Salt Lake City, UT

Dr. William Mitsch, Professor, Olentangy River Wetland Research Park, The Ohio State University, Columbus, OH

Dr. William Moomaw, Professor of International Environmental Policy and Director of the Center for International 

Environment and Resource Policy, The Fletcher School of Law and Diplomacy, Tufts University, Medford, MA

Dr. Arvin Mosier, Visiting Professor, Agricultural and Biological Engineering Department, University of Florida, Mount 

Pleasant, SC

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

Carolina – Chapel Hill, Morehead City, NC

Dr. Bryan Shaw, Commissioner, Texas Commission on Environmental Quality, Austin, TX

Mr. Paul Stacey, Director, Bureau of Water Management and Land Reuse, Planning and Standards Division, 

Connecticut Department of Environmental Protection, Hartford, CT

U.S. Environmental Protection Agency 

Science Advisory Board

Integrated Nitrogen Committee
iv

SCIENCE ADVISORY BOARD STAFF

Dr. Thomas Armitage, Designated Federal Officer (2009-present), U.S. Environmental Protection Agency, 

Washington, DC

Dr. Angela Nugent, Designated Federal Officer (2009), U.S. Environmental Protection Agency, Washington, DC

Ms. Kathleen White, Designated Federal Officer (2007-2008), U.S. Environmental Protection Agency, Washington, DC
v

CHAIR

Dr. Deborah L. Swackhamer, Professor and Charles M. Denny, Jr., Chair in Science, Technology and Public Policy and 

Co-Director of the Water Resources Center, Hubert H. Humphrey Institute of Public Affairs, University of Minnesota, St. 

Paul, MN

SAB MEMBERS

Dr. David T. Allen, Professor, Department of Chemical Engineering, University of Texas, Austin, TX

Dr. Claudia Benitez-Nelson, Full Professor and Director of the Marine Science Program, Department of Earth and 

Ocean Sciences, University of South Carolina, Columbia, SC

Dr. Timothy Buckley, Associate Professor and Chair, Division of Environmental Health Sciences, College of Public 

Health, The Ohio State University, Columbus, OH

Dr. Patricia Buffler, Professor of Epidemiology and Dean Emerita, Department of Epidemiology, School of Public 

Health, University of California, Berkeley, CA

Dr. Ingrid Burke, Director, Haub School and Ruckelshaus Institute of Environment and Natural Resources, University 

of Wyoming, Laramie, WY

Dr. Thomas Burke, Professor, Department of Health Policy and Management, Johns Hopkins Bloomberg School of 

Public Health, Johns Hopkins University, Baltimore, MD

Dr. Terry Daniel, Professor of Psychology and Natural Resources, Department of Psychology, School of Natural 

Resources, University of Arizona, Tucson, AZ

Dr. George Daston, Victor Mills Society Research Fellow, Product Safety and Regulatory Affairs, Procter & Gamble, 

Cincinnati, OH

Dr. Costel Denson, Managing Member, Costech Technologies, LLC, Newark, DE

Dr. Otto C. Doering III, Professor, Department of Agricultural Economics, Purdue University, W. Lafayette, IN

Dr. David A. Dzombak, Walter J. Blenko Sr. Professor of Environmental Engineering, Department of Civil and 

Environmental Engineering, College of Engineering, Carnegie Mellon University, Pittsburgh, PA

Dr. T. Taylor Eighmy, Vice President for Research, Office of the Vice President for Research, Texas Tech University, 

Lubbock, TX

Dr. Elaine Faustman, Professor and Director, Institute for Risk Analysis and Risk Communication, School of Public 

Health, University of Washington, Seattle, WA

Dr. John P. Giesy, Professor and Canada Research Chair, Veterinary Biomedical Sciences and Toxicology Centre, 

University of Saskatchewan, Saskatoon, Saskatchewan, Canada

Dr. Jeffrey K. Griffiths, Professor, Department of Public Health and Community Medicine, School of Medicine, Tufts 

University, Boston, MA

Dr. James K. Hammitt, Professor, Center for Risk Analysis, Harvard University, Boston, MA

Dr. Bernd Kahn, Professor Emeritus and Associate Director, Environmental Radiation Center, Georgia Institute of 

Technology, Atlanta, GA

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

Providence, RI

Dr. Madhu Khanna, Professor, Department of Agricultural and Consumer Economics, University of Illinois at 

Urbana-Champaign, Urbana, IL

Dr. Nancy K. Kim, Senior Executive, Health Research, Inc., Troy, NY

U.S. Environmental Protection Agency

Science Advisory Board
vi

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

Dr. Kai Lee, Program Officer, Conservation and Science Program, David & Lucile Packard Foundation, Los Altos, CA

Dr. Cecil Lue-Hing, President, Cecil Lue-Hing & Associates Inc., Burr Ridge, IL

Dr. Floyd Malveaux, Executive Director, Merck Childhood Asthma Network, Inc., Washington, DC

Dr. Lee D. McMullen, Water Resources Practice Leader, Snyder & Associates, Inc., Ankeny, IA

Dr. Judith L. Meyer, Professor Emeritus, Odum School of Ecology, University of Georgia, Lopez Island, WA

Dr. James R. Mihelcic, Professor, Civil and Environmental Engineering, State of Florida 21st Century World Class 

Scholar, University of South Florida, Tampa, FL

Dr. Jana Milford, Professor, Department of Mechanical Engineering, University of Colorado, Boulder, CO

Dr. Christine Moe, Eugene J. Gangarosa Professor, Hubert Department of Global Health, Rollins School of Public 

Health, Emory University, Atlanta, GA

Dr. Horace Moo-Young, Dean and Professor, College of Engineering, Computer Science, and Technology, California 

State University, Los Angeles, CA

Dr. Eileen Murphy, Grants Facilitator, Ernest Mario School of Pharmacy, Rutgers University, Piscataway, NJ

Dr. Duncan Patten, Research Professor, Hydroecology Research Program, Department of Land Resources and 

Environmental Sciences, Montana State University, Bozeman, MT

Dr. Stephen Polasky, Fesler-Lampert Professor of Ecological/Environmental Economics, Department of Applied 

Economics, University of Minnesota, St. Paul, MN

Dr. C. Arden Pope III, Professor, Department of Economics, Brigham Young University, Provo, UT

Dr. Stephen M. Roberts, Professor, Department of Physiological Sciences, Director, Center for Environmental and 

Human Toxicology, University of Florida, Gainesville, FL

Dr. Amanda Rodewald, Professor of Wildlife Ecology, School of Environment and Natural Resources, The Ohio State 

University, Columbus, OH

Dr. Jonathan M. Samet, Professor and Flora L. Thornton Chair, Department of Preventive Medicine, University of 

Southern California, Los Angeles, CA

Dr. James Sanders, Director and Professor, Skidaway Institute of Oceanography, 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, Philip E. Austin Professor of Economics, Department of Economics, University of Connecticut, 

Storrs, CT

Dr. Herman Taylor, Director, Principal Investigator, Jackson Heart Study, University of Mississippi Medical Center, 

Jackson, MS

Dr. Barton H. (Buzz) Thompson, Jr., Robert E. Paradise Professor of Natural Resources Law at the Stanford Law 

School and Perry L. McCarty Director, Woods Institute for the Environment, Stanford University, Stanford, CA

Dr. Paige Tolbert, Professor and Chair, Department of Environmental Health, Rollins School of Public Health, Emory 

University, Atlanta, GA

Dr. John Vena, Professor and Department Head, Department of Epidemiology and Biostatistics, College of Public 

Health, University of Georgia, Athens, GA

Dr. Thomas S. Wallsten, Professor and Chair, Department of Psychology, University of Maryland, College Park, MD

Dr. Robert Watts, Professor of Mechanical Engineering Emeritus, Tulane University, Annapolis, MD

Dr. R. Thomas Zoeller, Professor, Department of Biology, University of Massachusetts, Amherst, MA

SCIENCE ADVISORY BOARD STAFF

Dr. Angela Nugent, Designated Federal Officer, U.S. Environmental Protection Agency, Washington, DC
vii

Acknowledgements

The EPA Science Advisory Board Integrated Nitrogen Committee would like to acknowledge many individuals who 

provided their perspectives and insights for the Committee’s consideration in the development of this report.

Invited speakers provided perspectives at public meetings of the Committee. These speakers included: the following 

individuals from the U.S. EPA - Mr. Robert Bastian, Mr. Gilbert Castellanos, Mr. John Davies, Dr. Robin Dennis, Dr. 

Jonathan Garber, Dr. Richard Haeuber, Dr. Alan Hecht, Mr. Rudolph Kapichak, Mr. Gary Lear, Dr. Richard Linthurst, 

Dr. Rohit Mathur, Ms. Roberta Parry, and Mr. James Pendergast; the following individuals from the USDA - Dr. Roger 

Claassen, Dr. Stan Daberkow, Dr. Raymond Knighton, and Dr. Mark Walbridge; and the following individuals from other 

organizations - Mr. Craig Cox, Soil and Water Conservation Society; Dr. Jan Willem Erisman, Energy Research Centre 

of the Netherlands; Dr. Paul Fixen, International Plant Nutrition Institute; Dr. David McNaught, Environmental Defense 

Fund; Dr. John Miranowski, Iowa State University; Dr. Martin Petrovic, Cornell University; and Mr. John Sheehan, 

LiveFuels. 

The Committee would also like to thank the speakers and participants who attended the SAB Integrated Nitrogen 

Committee workshop held on October 20-22, 2008. These participants included: the following individuals from the 

U.S. EPA - Mr. Richard Batiuk, Dr. Jana Compton, Dr. Robin Dennis, Ms. Katie Flahive, Dr. Richard Haeuber, Ms. 

Chris Lewiciki, Dr. James Liebman, Ms. Sally Shaver, Ms. Roberta Parry, Mr. Randy Waite, Mr. John Wilson, Mr. Tom 

Wirth, and Mr. Marcus Zobrist; the following individuals from the USDA - Dr. Dan Jaynes, Dr. Ray Knighton, and Mr. 

Richard Swenson; and the following individuals from other organizations - Dr. Jan Willem Erisman, Energy Research 

Centre of the Netherlands; Mr. Albert Ettinger, Environmental Law and Policy Center; Mr. John Hardin, John Hardin 

& Son; Dr. Shelie Miller, University of Michigan; Dr. Catherine O’Connor, Metropolitan Water Reclamation District of 

Greater Chicago; Mr. Robin O’Malley, The Heinz Center; Mr. Richard Poirot, Vermont Agency of Natural Resources; 

Mr. John Quinn, Constellation Energy; Mr. Pat Rice, Nebraska Department of Environmental Quality; Dr. Joe Rudek, 

Environmental Defense Fund; Dr. Ted Russell, Georgia Institute of Technology; Ms. Mindy Selman, World Resources 

institute; Dr. Thomas Simpson, University of Maryland; Dr. Richard Smith, U.S. Geological Survey; Dr. Clifford Snyder, 

International Plant Nutrition Institute; Dr. Robert Summers, Maryland Department of the Environment; Dr. David 

Whitall, National Oceanic and Atmospheric Administration; Ms. Marcia Willhite, Illinois Environmental Protection 

Agency; and Mr. James Wurtz, Railway Equipment Company.

In addition, the Committee thanks the experts who provided independent review of the report in draft form: Dr. 

John Day, Louisiana State University; Dr. Elisabeth Holland, National Center for Atmospheric Research; Dr. Gregory 

McIsaac, University of Illinois; Dr. Jerry Melillo, Marine Biological Laboratory, Woods Hole, Massachusetts; Dr. Gyles 

Randall, University of Minnesota; Dr. James Schauer, University of Wisconsin-Madison; and Dr. Stuart Weiss, Creekside 

Center for Earth Observations. The reviewers provided many constructive comments and suggestions, but were not, 

however, asked to endorse the conclusions or recommendations in the report.
viii

NOTICE

This report has been written as part of the activities of the EPA Science Advisory Board (SAB), a public advisory 

group 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, nor does mention of trade names of commercial products constitute a 

recommendation for use. Reports of the SAB are posted on the EPA website at http://www.epa.gov/sab.
ix

List of Figures . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . xi

List of Tables . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . xiii

List of Boxes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . xiv

List of Chemical Abbreviations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . xv

List of Acronyms and Abbreviations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . xvi

Executive Summary . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ES-1

1. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3

 1.1. Overview of the Problem – Impacts of Excess Reactive Nitrogen (Nr) 

on Human Health and the Environment. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3

 1.2. The Nitrogen Cascade – Nr Loading, Cycling, and Exposure. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3

 1.3. EPA Activities to Manage Risks Posed by Nr. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6

 1.4. SAB Integrated Nitrogen Committee Study Objectives . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7

 1.5. Study Approach and Structure of the Report. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8

2. Sources, Transfer, and Transformation of Nr 

in Environmental Systems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11

 2.1. Nr Flux in the Environment. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11

 2.2. Sources of New Nr to the Environment. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13

	 2.2.1.	 Nr Formation and Losses to the Environment from Fossil Fuel Combustion. . . . . . . . . . . . . . . . . . . . . . . 14

	 2.2.2.	 Nr Inputs and Losses to the Environment from Crop Agriculture . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15

	 2.2.3.	 Nr Inputs and Losses from Animal Agriculture . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 25

	 2.2.4.	 Nr Inputs to Residential and Recreational Turf Systems. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 29

 2.3. Nr Transfer and Transformations in and Between Environmental Systems. . . . . . . . . . . . . . . . . . . . . . . . . . . 31

	 2.3.1.	 Input and Transfers of Nr in the United States. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 31

	 2.3.2.	 Storage of Nr Within Terrestrial Environmental Systems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 35

	 2.3.3.	 Areas of Uncertainty in Nr Transfer and Transformation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 38

3. Impacts of Nr on Aquatic, Atmospheric, and Terrestrial Ecosystems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 41

 3.1. Impacts on Drinking Water, Human Health, and Freshwater Biota. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 41

 3.2. Impacts of Airborne or Atmospherically Deposited Nr on Human Health 

and Ecosystems. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 42

 3.3. Impacts of Nr on Terrestrial Ecosystems. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 42

4. Metrics and Current Risk Reduction Strategies for Reactive Nitrogen . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 43

 4.1. Measurement of Nr in the Environment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 43

 4.2. Consideration of Nr Impacts in Risk Reduction Strategies. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 43

 4.3. Water Quality Regulation and Management. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 45

 4.4. Water Quality Monitoring and Assessment. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 49

 4.5. Clean Air Act and Air Quality Regulation and Management . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 50

 4.6. Thresholds for Excess Nr Effects on Terrestrial Ecosystems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 52

 4.7. Comments on Nr Critical Loads. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 52

 4.8. Tradeoffs of Nr Impacts in Risk Reduction Strategies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 53

 4.9. Interactions of the N Cascade and Climate. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 56

Table of Contents
x

5. Integrated Risk Reduction Strategies for Nr . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 59

 5.1. Importance of Integrated Risk Reduction Strategies. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 59

 5.2. Control Strategies for Nr. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 59

 5.3. Management Strategies for Nr in the Environment. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 59

	 5.3.1.	 Command-and-control. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 59

	 5.3.2.	 Direct Allocation of Federal Funds for Conservation Programs. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 60

	 5.3.3.	 Market Based Instruments for Pollution Control . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 61

	 5.3.4.	 Biophysical and Technical Controls (control points) on Transfer

and Transformations of Nr in and Between Environmental Systems. . . . . . . . . . . . . . . . . . . . . . . . . . . . . 62

6. SAB Recommendations for Nr Data Collection, Risk Management, and Research . . . . . . . . . . . . . . . . . . . . 73

 6.1. Need for Comprehensive Monitoring of Nr. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 73

 6.2. Overarching Recommendations. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 74

 6.3. Near-term Management Goals. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 75

 6.4. Summary of Specific Findings and Recommendations 

Corresponding to the Four Study Objectives. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 79

 6.5. Conclusions and Observations. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 85

Appendix A: Nitrogen Deposition From the Atmosphere to the Earth’s Surface . . . . . . . . . . . . . . . . . . . . . . . . 89

Appendix B: Sources and Cycling of Nr Input 

into Terrestrial Systems in the United States . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 95

Appendix C: Water Quality Trading in the Illinois River Basin . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 99

Appendix D: Management of Nr Measures Based 

on the Concept of Critical Loads . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 103

Appendix E: Technical Annexes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 107

Appendix F: Recent Major EPA Mobile Source Rules to Control NOx . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 109

Appendix G: Impacts of Reactive Nitrogen on Aquatic Systems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 111

Appendix H: Nr Saturation and Ecosystem Function . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 121

References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 123
xi

List of Figures

Figure ES-1: The nitrogen cascade. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ES-3

Figure ES-2: Sources of reactive nitrogen (Nr) introduced into the United States in 2002 (Tg N/yr). . . . . . . . . . . ES-5

Figure 1: The nitrogen cascade . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5

Figure 2: Sources of reactive nitrogen (Nr) introduced into the United States in 2002 (Tg N/yr). . . . . . . . . . . . . . . . 11

Figure 3: U.S. NOx emission trends, 1970-2006. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14

Figure 4: Percent reductions in NOx emissions, 1990-2002, from different sources 

(off-road, on-road, power generation, etc.) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14

Figure 5: Mobile source NOx emission inventories. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15

Figure 6: Fertilizer consumption in the United States, 1960 to 2006 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17

Figure 7: Trends in corn grain produced per unit of applied fertilizer N (NFUE) in the United States. . . . . . . . . . . . 19

Figure 8: Synthetic fertilizer and livestock manure N used as fertilizer in Denmark 1985-2003. . . . . . . . . . . . . . . . 21

Figure 9: Protein content of cereal grain in Denmark. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 21

Figure 10: Meat production from 1970 to 2006. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 25

Figure 11: Milk production from 1970 to 2006. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 25

Figure 12: U.S. inventory of mature dairy cows and milk production per cow from 1970 to 2006. . . . . . . . . . . . . . . 26

Figure 13: Number of animal operations in the United States from 1970 to 2006. . . . . . . . . . . . . . . . . . . . . . . . . . . . 26

Figure 14: Nr input and loss from 16 watersheds in the northeast United States. . . . . . . . . . . . . . . . . . . . . . . . . . . . . 39

Figure 15: Relative importance of all reactive nitrogen sources released into atmospheric, terrestrial, 

and freshwater media within the Chesapeake Bay Watershed utilizing four different metrics. . . . . . . . . . 46

Figure 16: Quantified damage costs (including health impacts) relative to tonnes of reactive nitrogen 

in Chesapeake Bay Watershed. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 46

Figure 17: Diagram of the nitrification and denitrification processes. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 54

Figure 18: Combined carbon and nitrogen global cycles . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 55

Figure 19: Comparisons between Global Warming Potential (GWP) and eutrophication impact 

categories for various bioproducts. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 56

Figure 20: Probability of given discharge level for nitrate in the watersheds of eastern Iowa . . . . . . . . . . . . . . . . . . . 57

Figure 21: Relative nitrogen discharge (lbs/day) from 79 POTWs. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 64

Figure 22: Trading ratios for municipalities in Connecticut. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 65

Figure 23: The likely impact of research investment in increasing N fertilizer use efficiency. . . . . . . . . . . . . . . . . . . 68

Figure A-1: Percent change in relative contribution of oxidized (NO3

-

) and reduced (NH4

+) 

nitrogen wet deposition from 1994 to 2006. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 89

Figure A-2: Trend in reported wet deposition of NH4

+ and NO3-

 for the 48 contiguous states. . . . . . . . . . . . . . . . . . 90

Figure A-3: Annual NH4

+, NO3-

, and total inorganic N deposition for the year 2007 showing 

spatial patterns of deposition . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 90

Figure A-4: CMAQ annual average (wet plus dry and oxidized plus reduced) nitrogen deposition 

(in kg-N/ha/yr) across the United States . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 93

Figure C-1: Distribution of municipal discharge, and industrial dischargers in the 

Illinois River Watershed . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 100

Figure C-2: Distribution of total nitrogen emissions by sub-watershed . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 100

Figure C-3: Potential land availability in the 100-year flood zone for nutrient farming in each 

sub-watershed in the Illinois River Watershed . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 100

Figure C-4: Spring available total nitrogen load by sub-watershed. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 100
xii

Figure C-5: Spring marginal cost (price) by watershed. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 101

Figure C-6: Unrestricted spring credit sales (tons/month) by sub-watershed. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 101 

Figure E-1: Total Nr yields (kg/ha/yr) in large rivers of the U.S. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 108

Figure G-1: U.S. Population (1995) supplied by domestic drinking water wells. . . . . . . . . . . . . . . . . . . . . . . . . . . . 112

Figure G-2: Nitrate concentrations in U.S. domestic drinking water wells. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 113
xiii

List of Tables

Table ES-1: Examples of impacts of excess reactive nitrogen on human health and environment. . . . . . . . . . . . . ES-2

Table 1: Nr fluxes for the United States, Tg N in 2002. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12

Table 2: Examples of multiple sources from states with high NOx emissions based on 2001 data; 

and tons of NOx as NO2. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15

Table 3: Types and amount of nitrogen fertilizers used in the United States in 2002. . . . . . . . . . . . . . . . . . . . . . . . . . 18

Table 4: Estimates of nitrogen input from biological nitrogen fixation from major legume crops, 

hay, and pasture. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22

Table 5: N2O emissions in the United States, 2002. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23

Table 6: Livestock N excretion per kg production (g/kg) and per total United States (Tg/yr). . . . . . . . . . . . . . . . . . . 28

Table 7: Manure production from animal husbandry in the continental United States, Tg N per year 2002 . . . . . . . . 28

Table 8: Fate of livestock manure nitrogen (Tg N). . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 29

Table 9: Estimate of fertilizer N used on turf grass in the United States in the year 2000, based on 

total area of 12.6 million ha. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 30

Table 10: Net annual change in continental U.S. croplands soil N and C, forest C and N, and 

grasslands C and N in 2002. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .37

Table 11: Ecosystem service and corresponding function categories. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 44

Table 12: Marginal abatement cost per tonne of Nr by source . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 47

Table 13: Federal primary ambient air quality standards that involve Nr, effective February 2010. . . . . . . . . . . . . . . 50

Table 14: Advantages and limitations of various approaches to Nr control in forestry and agriculture. . . . . . . . . . . . 60

Table 15: Summary of market-based instruments for pollution control with conceptual examples. . . . . . . . . . . . . . . 63

Table 16: Performance of the Nitrogen Credit Exchange . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 64

Table 17: Estimates for potential decreases in NH3 emissions from livestock manure in the United States. . . . . . . . 78

Table A-1: Annual wet deposition of reduced (NH4

+), oxidized (NO3-

), and total N to the 

48 contiguous states. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 89

Table A-2: Deposition of N to the eastern United States in units of kg N/ha/yr. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 91

Table A-3: Results from CMAQ for total deposition in 2002 to the 48 contiguous states of oxidized 

and reduced N . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 92

Table B-1: Sources of reactive N into terrestrial systems in the United States in 2002 

(from Table 1 data sources; in Tg N/yr). . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 95

Table B-2: Nr input and flows (Tg N/yr) in the terrestrial portion of the Nitrogen Cascade 

within the continental United States in 2002. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 96

Table C-1: Nutrient farm market parameters under three trading scenarios. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 101

Table D-1: Summary of the effects of excess Nr on human health in relation to metrics, current 

international regulations and conventions, and the link to the nitrogen cascade. . . . . . . . . . . . . . . . . . . 104

Table D-2: Summary of the effects of excess Nr on ecosystems related to currently used metrics, 

the existence of European regulatory values, and the link to the nitrogen cascade. . . . . . . . . . . . . . . . . 105

Table D-3: Summary of the effects of excess N on other societal values in relation to metrics and 

regulatory values in current international regulations and conventions and the link to 

the nitrogen cascade. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 106

Table G-1: Estuaries with nitrogen management plans or TMDLs and percent nitrogen load reduction targets. . . . 118
xiv

List of Boxes

Box 1: Hypoxia in the Gulf of Mexico. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 35

Box 2: Economic impact and metrics for Chesapeake Bay and its watershed . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 46

Box 3: The impact of climate change on agricultural discharge of reactive nitrogen. . . . . . . . . . . . . . . . . . . . . . . . . 57

Box 4: Water quality trading to meet the Long Island Sound wasteload allocation in Connecticut. . . . . . . . . . . . . . . 63

Box G-1: The national criterion for ammonia in fresh water . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 114

Box G-2: Long Island Sound total maximum daily load: focus on reactive nitrogen. . . . . . . . . . . . . . . . . . . . . . . . 119
xv

List of Chemical Abbreviations

C Carbon

CFC Chlorofluorocarbon 

DIN Dissolved inorganic nitrogen

DO Dissolved oxygen

Fe Iron

H Hydrogen

HNO3 Nitric acid

HONO Nitrous acid

N Nitrogen 

N2 Diatomic (molecular) nitrogen

N2O Nitrous oxide, 

N2O5 Dinitrogen pentoxide (nitric acid anhydride)

NH3 Ammonia

NH4

+ Ammonium

NHx NH3 NH4

+

NO Nitric oxide

NO2 Nitrogen dioxide

NO3

- Nitrate ion

NO3 Nitrate radical

Norg Organic nitrogen

NOx Nitrogen oxides (NO + NO2)

NOy Total reactive oxidized nitrogen

(NO, NO2, NO3, 2xN2O5, HONO, HNO3, NO3

-

, PAN and other organo-nitrates, RONO2)

Nr Reactive nitrogen 

O2 Oxygen

OH Hydroxyl radical

P Phosphorus

PAN Peroxy acetyl nitrate

PM Particulate matter

PM2.5 Particulate matter less than 2.5 microns in diameter

PM10 Particulate matter less than 10 microns in diameter

RONO2 Organic nitrates

Si Silicon

SO2 Sulfur dioxide

SO4

2- Sulfate

TAN Total ammonical nitrogen
xvi

List of Acronyms and Abbreviations

AAPFCO Association of American Plant Food Control Officials

AARA American Reinvestment and Recovery Act

AIRMON Atmospheric and Integrated Research Monitoring Network

AOB Ammonia oxidizing bacteria

BL Boundary layer

BMP Best management practice

BNF Biological nitrogen fixation

BNR Biological nutrient (or nitrogen) removal

CAA Clean Air Act

CAFO Concentrated animal feeding operation

CAIR Clean Air Interstate Rule 

CALM Consolidated Assessment and Listing Methodology

CAST Council for Agricultural Science and Technology

CASTNET Clean Air Standards and Trends Network 

C-BNF Cultivation-induced biological nitrogen fixation 

CCC Criterion continuous concentration

CFC Chlorofluorocarbon 

CFR Code of Federal Regulations

CL Critical load (threshold of Nr loading at which negative impacts have been documented)

CLAD Critical Loads Ad-Hoc Committee

CMAQ Community multiscale air quality

CMC Criterion maximum concentration

CRP Conservation Reserve Program

CSO Combined sewer overflow

CTM Chemical Transport Models

CWA Clean Water Act

CWSRF Clean Water State Revolving Fund (construction grants program under the Clean Water Act)

DOE U.S. Department of Energy

DOT U.S. Department of Transportation 

EGU Electricity generating units

EFD Essential Facilities Doctrine

EGR Exhaust gas recirculation

EISA Energy Independence and Security Act

EPA United States Environmental Protection Agency

EQIP Environmental Quality Incentives Program

EU European Union

FAO Food and Agricultural Organization of the United Nations

FAOSTAT Food and Agricultural Organization Statistical Database

FGR Flue-gas recirculation
xvii

ha Hectare

GHG Greenhouse gas

GPS Geographic Positioning System

HAB Harmful algal bloom

IPCC Intergovernmental Panel on Climate Change

ISA Integrated Science Assessments

ITQ Individual transferable quota

kg Kilogram

L Liter

LA Load allocation

LCA Life cycle analysis

LISS Long Island Sound Study

MCL Maximum contaminant level

MCLG Maximum contaminant level goal

mg Milligrams

MGD Million gallons per day

MJ Megajoule (one million joules)

Mmt Million metric tons

MT metric tons

MOM Mississippi-Ohio-Missouri

MRB Mississippi River Basin

MS4 Municipal separate storm sewer system

NAAQS National Ambient Air Quality Standards

NADP National Atmospheric Deposition Program

NASS National Agricultural Statistics Service Information 

NCA National Coastal Assessment

NCE Nitrogen Credit Exchange 

NCCR National Coastal Condition Report

NEEA National Estuarine Eutrophication Assessment

NESCAUM Northeast States for Coordinated Air Use Management

NFUE Nitrogen fertilizer use efficiency. Calculated as the ratio of grain yield to 

the quantity of applied N fertilizer (kg grain/kg applied N).

NMP Nutrient management plan

NOAA National Oceanic and Atmospheric Administration

NPS Nonpoint source

NRC National Research Council

NRCS Natural Resources Conservation Service

NRD Natural Resource District

NRI National Resources Inventory

NTN National Trends Network
xviii

NUE Nitrogen use efficiency. NUE is defined as the kg grain produced 

per kg of total N used by the crop, where total N includes N from fertilizer, 

biological N fixation, and soil organic matter mineralization

OTAG Ozone Transport Assessment Group

OTC Ozone Transport Commission

PE Physiological efficiency (physiological efficiency with which the N 

taken up by the crop is used to produce economic yield such as grain or fruit, 

quantified by kg increase in economic yield per kg of N accumulation in above ground crop biomass) 

PFP Partial factor productivity

POTW Publicly owned treatment works

PSD Prevention of significant deterioration

RE Recovery efficiency (kg N uptake per kg N applied)

SAV Submerged aquatic vegetation

SNCR Selective non-catalytic reduction

SCR Selective catalytic reduction

SIP State Implementation Plan

SOM Soil organic matter

SPARROW Spatially Referenced Regressions on Watershed Attributes Model 

STP Sewage treatment plant

SW Storm water

SWAT Storm Water Assessment Tool

SWPPP Stormwater Pollution Prevention Plan

T Temperature 

Tg Teragram (million metric tons or 1012 grams)

TMDL Total maximum daily load

TN Total nitrogen

UFTRS Uniform Fertilizer Tonnage Reporting System

UNECE United Nations Economic Commission for Europe

USDA U.S. Department of Agriculture

USGS U.S. Geological Survey

USEPA United States Environmental Protection Agency

WHO World Health Organization

WLA Wasteload allocation

WPCA Water pollution control authorities 

WRI World Resources Institute

WRP Wetland Reserve Program

WSA Wadeable Stream Assessment
ES-1

Introduction

Nitrogen (N) is an integral component of all proteins, 

which are the basic building blocks of life and catalysts 

for life-sustaining reactions in organisms. Reactive 

nitrogen (Nr), in contrast to non-reactive gaseous N2, 

includes all biologically active, chemically reactive, and 

radiatively active nitrogen compounds in the atmosphere 

and biosphere of the earth.1 Without an adequate supply 

of N in any organism’s diet, it can’t survive. Ironically, 

bioavailable N for nutrition is in short supply and indeed 

the productivity of most of the world’s ecosystems is 

often limited by the availability of N. This is certainly 

the situation with food production. Without the creation 

of N fertilizer by an industrial process (the Haber-Bosch 

process) and the increased cultivation of leguminous 

crops, the world could not support the current human 

population or its projected increase.

As further discussed in Chapter 2 of this report, 

increased anthropogenic input of Nr to the environment 

has contributed to large increases in the mass flux of 

nitrogen via the nitrogen cycle. Anthropogenic sources 

of N now provide enough N, on average, to grow food 

for the world’s peoples. However, a major consequence 

of this nearly inexhaustible supply is that most N used 

in food production, and all of the new Nr produced by 

fossil fuel combustion, is lost to the environment where 

it circulates through the earth’s atmosphere, hydrosphere, 

geosphere, and biosphere. During this circulation, Nr 

contributes to a wide variety of consequences, which 

are magnified with time as Nr moves through the 

environment. 

Impacts of reactive nitrogen on human 

health and the environment 

Anthropogenic creation of Nr provides essential 

benefits for humans – first and foremost in meeting 

human dietary needs. A large fraction of the human 

population of the earth could not be sustained if synthetic 

nitrogen fertilizers did not significantly augment food 

production. Essentially all of the Nr created by human 

activities, however, is released to the environment, often 

with unintended negative consequences. As summarized 

in Table ES-1, it contributes to a number of adverse 

public health and environmental effects, including 

photochemical smog, decreased atmospheric visibility, 

acidification of terrestrial and aquatic ecosystems, 

eutrophication of coastal waters (i.e., harmful algal 

blooms, hypoxia), drinking water concerns, freshwater 

Nr imbalances, greenhouse gas emissions and subsequent 

climate change, and stratospheric ozone depletion. 

In light of the magnitude of the human alteration of the 

nitrogen cycle, and the resulting negative consequences 

on humans and ecosystems, the National Academy of 

Engineering has identified management of the nitrogen as 

one of the “grand challenges” facing this country.2

Nr effects are manifest as direct declines in both 

human health (e.g., respiratory and cardiac diseases) and 

ecosystem health (e.g., coastal eutrophication and loss in 

biodiversity). In addition, there are indirect declines in 

human health because the negative impacts on ecosystems 

will diminish the services that those ecosystems provide 

people. The effects are often magnified because the 

same atom of nitrogen can cause multiple effects in 

the atmosphere, in terrestrial ecosystems, in freshwater 

and marine systems, and on human health. We call this 

sequence of effects the nitrogen cascade. 

The nitrogen cascade 

The nitrogen cascade has three dimensions: 

biogeochemical, alterations in the environment, and 

human and ecosystem consequences. 

The “biogeochemical” dimension of the nitrogen 

cascade involves: Nr creation from N2 as a consequence 

of chemical, food, and energy production; Nr use in food 

and chemical production; Nr losses to the environment; 

changes in Nr species residence times in environmental 

reservoirs; Nr transfers among reservoirs; and Nr 

conversion back to N2. Alterations to the environment 

then result from increased Nr levels in the environment. 

These alterations have negative consequences for 

ecosystem and human health at local, regional, national, 

and global scales. Because nitrogen is a critical resource 

and also a contributor to many of the environmental 

concerns facing the U.S. today, it is imperative to 

understand how human action has altered N cycling in the 

U.S., and the consequences of those alterations on people 

and ecosystems. The overarching question is, how do 

we protect and sustain ecosystems that provide multiple 

benefits to society while also providing the interconnected 

material, food and energy required by society?

Executive Summary

1 Reactive nitrogen (Nr) includes inorganic chemically reduced forms of N (NHx) [e.g., ammonia (NH3) and ammonium ion (NH4+)], inorganic 

chemically oxidized forms of N [e.g., nitrogen oxides (NOx), nitric acid (HNO3), nitrous oxide (N2O), N2O5, HONO, peroxy acetyl compounds such 

as peroxyacytyl nitrate (PAN), and nitrate ion (NO3

-

)], as well as organic compounds (e.g., urea, amines, amino acids, and proteins).

2 National Academy of Engineering Grand Challenges (http://www.engineeringchallenges.org/cms/challenges.aspx)
ES-2

Impact Cause Location Metric Source Reference

AIR

Visibility decrease Fine particulate 

matter

National Parks and 

wilderness areas visibility impairment

NOy and NHx from 

fossil fuels and 

agriculture

Malm et al., 

2004; U.S. 

EPA Clean Air 

Scientific Advisory 

Committee, 

2004 EPACASAC-09-010

LAND - ECOSYSTEMS

Biodiversity loss Nitrogen deposition

Grasslands and 

forests in the United 

States receiving N 

deposition in excess 

of critical load

Decrease in species 

richness of grasslands and forests

Utilities, traffic, and 

animal agriculture

Bobbink et al., 

2010; Fenn et al., 

2003.

Forest decline Ozone and acid 

deposition

Eastern and Western United States

Decreased timber 

growth; increased 

susceptibility to 

disease and pests

Utilities, traffic, and 

animal agriculture

Johnson & 

Siccama, 1983; 

MacKenzie & 

El-Ashry, 1990

LAND - AGRICULTURE

Crop yield loss Ozone Eastern and Western United States $ 2-5 billion/year Utilities & traffic Heck et al., 1984

WATER

Acidification of 

surface waters; loss 

of biodiversity

Acidification of 

soils, streams and 

lakes is caused 

by atmospheric 

deposition of 

sulfur, HNO3, NH3

and ammonium 

compounds. 

Primarily mountainous regions of the 

United States

Out of 1,000 lakes 

and thousands of 

miles of streams 

in the Eastern 

United States 

surveyed, 75% of 

the lakes and 50% 

of the streams were 

acidified by acid 

deposition

Fossil fuel combustion and agriculture

U.S. EPA, 2008a 

http://www.epa.

gov/acidrain

Hypoxia of coastal 

waters

Excess nutrient 

loading, eutrophication, variable 

freshwater runoff 

Gulf of Mexico, 

other estuarine and 

coastal waters

Benthic finfish/shellfish habitat loss, fish 

kills, sulfide toxicity, 

costs >$50 million 

annually

N, P from energy 

and food production

Bricker et al., 

1999; Verity et 

al., 2006; U.S. 

EPA SAB, 2007; 

Rabalais et al., 

1999; Mitsch et 

al., 2001

Harmful Algal 

Blooms

Excessive nutrient 

loading, climatic 

variability

Inland and coastal 

waters

Fish kills, losses 

of drinking and 

recreational waters 

costs >$100 million 

annually

Excess nutrient 

(N & P) loading 

Paerl, 1988; 

ECOHAB, 1995; 

NRC, 2000

HUMAN AND ENVIRONMENTAL HEALTH DAMAGES

Human mortality PM2.5, O3 and 

related toxins.

U.S. urban and 

nearby areas

 Pollution related 

deaths estimated 

at 28,000-55,000 

per year (a range 

of cardiovascular 

and respiratory 

system effects are 

associated with this 

pollution).

NOy and NHx from 

fossil fuels and 

agriculture

Mokdad et al., 

2004; Ezzati et 

al., 2004.

Total damage to 

public health and 

environment

NOx into air Chesapeake Bay 

Watershed

$3.4 Billion; 

200,000 MT Mobile sources

Moomaw and 

Birch, 2005; Birch 

et al., 2011

Total damage to 

public health and 

environment

NHx and nitrate into 

air and water

Chesapeake Bay 

Watershed

$1.5 Billion; 

400,000 MT Agriculture Moomaw and Birch, 2005

Table ES-1: Examples of impacts of excess reactive nitrogen on human health and environment
ES-3

Nr inputs to the nation and the world have been 

increasing, largely due to human activities associated 

with food production and fossil fuel combustion. Despite 

the obvious benefits of a plentiful supply of food and 

energy, the adverse consequences associated with the 

accumulation of Nr in the environment are large, with 

implications for human health and the environment. 

The greater the inputs of Nr to the landscape, the 

greater the potential for negative effects caused by 

greenhouse gas (GHG) production, ground level ozone, 

acid deposition, and Nr overload that can contribute 

to climate change, degradation of soils and vegetation, 

acidification of streams, lakes and rivers, estuarine and 

coastal eutrophication, hypoxia, and habitat loss.

The growing nature of the Nr problem, and the adverse 

and intertwined consequences associated with Nr inputs 

to air, land, and water as exhibited in the N cascade 

underscore the need for researchers and managers to 

explore integrated strategies that minimize N inputs, 

maximize its use efficiency, promote Nr removal 

processes, and protect humans and natural resources. 

The concept of the nitrogen cascade highlights that 

once a new Nr molecule is created, it can, in sequence, 

travel throughout the environment contributing to major 

environmental problems (Galloway et al., 2003). The 

adaptation of the cascade in Figure ES-1 was developed 

by the SAB Integrated Nitrogen Committee (INC) 

to provide a context for considering nitrogen-related 

issues and ecosystem effects in the U.S. To consider 

the cascading effects of Nr in the U.S., we examined 

the various atmospheric, terrestrial, and aquatic 

environmental systems where Nr is stored, and the 

magnitudes of the various flows of N to, from, and within 

them. The nitrogen cascade concept implies the cycling 

of Nr among these systems. The process of denitrification 

is the only mechanism by which Nr is converted to 

chemically inert N2, “closing” the continuous cycle 

(Figure ES-1 shows only flows of reactive nitrogen, not 

N2). Denitrification can occur in any of the indicated 

reservoirs except the atmosphere.

The “new” N box in the Nitrogen Cascade depicts 

the two primary anthropogenic sources by which Nr 

originates – energy production and food production –

and where Nr from these sources enters ecosystems. 

Energy production includes both fossil fuel and biofuel 

combustion. Food production includes N fertilizer 

produced in the U.S., cultivation-induced biological N

(C-BNF) in the U.S., production of animals and crops 

in the U.S. for human consumption, and imports of 

N-containing fertilizer, grain and meat to the U.S. 

Figure ES-1: The nitrogen cascade

The Nitrogen Cascade Atmospheric

Terrestrial

Aquatic

Energy Production

& combustion of fossil fuels

New Nitrogen

Food Production

& creation of synthetic fertilizers

NOy, NHx, Norg

NOy, NHx, Norg

stratosphere

ozone troposphere

ozone depletion

greenhouse gases

particulate matter

crops animals

soils

people

soils

landscape

forests grasslands

soils

Agricultural Populated

Vegetated

Groundwater

Oceans

Coastal Bays

& Estuaries

Surface Water

& Wetlands

NOy, NHx, Norg

N2O

N2O

NOx

NH3

Norg

biogeochemical cycling

ecosystem productivity

acidification

eutrophication

NOx

NHx
ES-4

The atmospheric system box in the Figure ES-1 

indicates that tropospheric concentrations of both ozone 

and particulate matter are increased due to emissions 

of nitrogen oxides3 (NOx) to the atmosphere. The ovals 

illustrate that the increase in N2O concentrations, in turn, 

contribute to the greenhouse effect in the troposphere and 

to ozone depletion in the stratosphere. Except for N2O, 

there is limited Nr storage in the atmosphere. Losses of 

Nr from the atmospheric system include total oxidized 

nitrogen4 (NOy), reduced nitrogen5 (NHx), and organic 

nitrogen (Norg) deposition to terrestrial and aquatic 

ecosystems of the earth’s surface. There is little potential 

for conversion of Nr to N2 via denitrification in air. 

However, once airborne deposition of Nr occurs it will be 

subject to denitrification pathways via soil and water.

The terrestrial system box in the Figure ES-1 depicts 

that Nr enters agricultural lands via food production 

and is introduced to the entire terrestrial landscape via 

atmospheric deposition. Within agricultural regions there 

is cycling among soils, crops and animals, and then a 

transfer of Nr as food to populated regions, from which 

there are Nr losses to the environment (e.g., sewage, 

landfills). The ovals showing ecosystem productivity 

and biogeochemical cycling reflect that Nr is actively 

transported and transformed within the terrestrial system, 

and that as a consequence there are significant impacts 

on ecosystem productivity due to fertilization and 

acidification, often with resulting losses of biodiversity. 

There is ample opportunity for Nr storage in both biomass 

and soils. Losses of Nr from this system occur by leaching 

and runoff of NOy, NHx and Norg to aquatic ecosystems 

and by emissions to the atmospheric system as NOx, NH3, 

Norg, and N2O. There is potential for conversion of Nr to 

N2 via denitrification in the terrestrial system. 

The aquatic system box in the Figure ES-1 shows that 

Nr is introduced via leaching and runoff from terrestrial 

ecosystems and via deposition from atmospheric 

ecosystems. Connected with the hydrological cycle, 

there are Nr fluxes downstream with ultimate transport 

to coastal systems. Within the aquatic system, the 

ovals highlight two significant impacts of waterborne 

Nr acidification of freshwaters and eutrophication of 

fresh and coastal waters. Except for Nr accumulation in 

groundwater reservoirs, there is limited Nr storage within 

the hydrosphere. Losses of Nr from the aquatic system are 

primarily via N2O emissions to the atmospheric system. 

There is a very large potential for conversion of Nr to N2

via denitrification in water and wetlands. 

NOy, NHx and N2O are all components of Nr, but 

a fundamental difference is that the NOy and NHx are 

rapidly transferred from the atmosphere to receiving 

ecosystems due to a short atmospheric residence time 

(≤ 10 days) where they continue to contribute to the 

N cascade. Because of its longer residence time (~100 

years) however, N2O remains in the troposphere where it 

contributes to climate change, until it is transferred to the 

stratosphere, where it contributes to ozone depletion.

Trends in N inputs to the United States 

In 2002, humans introduced 29 teragrams (Tg) of 

newly formed reactive N into the U.S. through HaberBosch process production of fertilizers and industrial 

Nr, cultivation-induced biological nitrogen fixation (i.e., 

conversion of N2 to NH3 by microorganisms associated 

with some cultivated crops, for example, legumes), and 

fossil fuel combustion (Figure ES-2). By definition, prior 

to human presence in the U.S., there was no introduced 

anthropogenic Nr. Prior to 1900, no Haber-Bosch Nr 

was introduced, fossil fuel combustion introduced very 

small amounts relative to today, and cultivation-induced 

biological nitrogen fixation created approximately 2 

Tg N. Thus, between 1900 and 2002, the amount of Nr 

introduced to the U.S. has increased by approximately 

10-fold. 

Nitrogen inputs to the United States

The EPA Science Advisory Board (SAB) Integrated 

Nitrogen Committee (“Committee”) evaluated nitrogen 

inputs to the U.S. in 2002. At the global scale, human 

activities produced approximately twice as much Nr as 

did natural processes. In the U.S., however, the amount of 

Nr produced by human activities was approximately five 

times larger than natural processes. As shown in Figure 

ES-2, natural ecosystems in the U.S. introduce about 6.4 

Tg of Nr as N per year (Tg N/yr). In contrast, human 

activities introduce about 28.5 Tg N/yr.

Chapter 2 of this report discusses sources, transfer, 

and transformation of Nr. Supporting references for 

the information presented on this topic are presented in 

Chapter 2. The largest single source of Nr in the U.S. is 

the Haber-Bosch process, which introduces about 15.2 

Tg N/yr: 9.4 Tg N/yr from domestic Nr production and 

5.8 Tg N/yr from imports of Nr in fertilizers. The 15.2 

Tg N/yr of anthropogenic Nr is used in three ways: 9.9 

Tg N/yr is used to produce agricultural crops; 1.1 Tg N/

yr is applied to turf grasses; and 4.2 Tg N/yr is used by 

industry for production of nylon, refrigerants, explosives 

and other commercial products.

The second largest source of Nr introduced into the 

U.S. is enhancement of biological nitrogen fixation (BNF) 

by cultivation of legumes like soybeans and alfalfa that 

have nitrogen-fixing symbionts, or by crops like rice 

that have nitrogen-fixing bacteria in their rhizosphere. 

These Nr fixing crops introduce about 7.7 Tg N/yr. A 

3 NOx (oxides of nitrogen) includes NO + NO2 4 NOy (total oxidized nitrogen) includes NO, NO2, NO3, N2O5, HONO, HNO3, NO3–, PAN and other organo-nitrates, RONO2 5 NHx (reduced nitrogen) includes NH3 + NH4
ES-5

small amount of additional Nr is also imported in grain 

and meat products; in 2002 this source of added Nr was 

approximately 0.2 Tg N/yr (not shown in Figure ES-2).

Fossil fuel combustion is the third largest source of 

new Nr. It introduces approximately 5.7 Tg N/yr into 

the environment (almost entirely as NOx), that is, 3.8 Tg 

N/yr from transportation sources and 1.9 Tg N/yr from 

stationary sources such as electric utilities, industrial 

boilers and from certain industrial processes.

In summary, agriculture and domestic use of fertilizers 

to produce food, feed, and fiber (including bioenergy 

and BNF) and combustion of fossil fuels are the largest 

sources of new Nr released into the environment in the 

U.S. The percentage distribution of Nr released to the U.S. 

environment from human activities in 2002 was: about 

65% from agricultural sources (including BNF and turf 

production), about 20% from fossil fuel sources, and about 

15% from industrial sources (Figure ES-2).

Distribution of reactive nitrogen through 

the environment

Once introduced into the U.S., Nr compounds are 

distributed via the atmosphere, hydrosphere, geosphere, 

biosphere, and commerce. Distribution in the atmosphere 

begins with NOx, NH3, and N2O. NOx and NH3 (and their 

reaction products) are distributed on a scale of hundreds 

to thousands of kilometers within the U.S. boundaries, 

and also distributed to downwind countries and oceans. 

Due to its long lifetime (approximately 100 years) in the 

atmosphere, N2O accumulates in the U.S. atmosphere 

4.2 6.4

3.8

1.9

7.7

10.9

Transportation

Stationary

Cultivation BNF

Haber Bosch N Fertilizer

Nonfertilizer Haber Bosch N 

BNF

Agriculture

Natural

Fossil Fuel

Industry

Figure ES-2: Sources of reactive nitrogen (Nr) introduced into the United States in 2002 (Tg N/yr).

Figure ES-2 explanatory notes:

 Numerical units: teragram of reactive nitrogen (Nr) per year (Tg N/yr)

 Natural BNF: biological nitrogen fixation in natural grasslands, rangelands, and forests,

 Fossil Fuel-Transportation: combustion in vehicles, trains, airplanes, ships and off-road construction equipment.

 Fossil Fuel-Stationary: combustion of fossil fuels in power plants and industrial boilers.

Agriculture-cultivation BNF: agricultural augmentation of biological nitrogen fixation – for example by planting of 

nitrogen fixing legumes.

Agriculture-Haber Bosch N fertilizer: agricultural (including turf production) use of synthetic nitrogen fertilizers 

produced by the Haber Bosch process for converting gaseous N2 to Nr. 

 Industry-Haber Bosch N: Industrial sources of Nr produced by the Haber-Bosch process.

 Figure ES-2 documents only the introduction of new Nr in the United States, and not the transfers of existing Nr among 

systems (e.g., Nr in manure).
ES-6

and is also dispersed throughout the global atmosphere. 

All ecosystems in the conterminous U.S. receive 

anthropogenic Nr from the atmosphere and for many 

ecosystems it is their primary, albeit unintended, source of 

Nr. Once deposited, Nr can be stored in soils and biomass 

and widely distributed via the stream-river continuum to 

inland and coastal waters. Some of the Nr is converted to 

N2O or denitrified to N2, primarily in aquatic ecosystems, 

including wetlands. Commerce is a major mechanism that 

transfers Nr from one place to another in the U.S.; most of 

the Nr that is used to produce food (e.g., fertilizer) and in 

food products crosses state boundaries via roads, railroads 

and the air.

Putting values to this distribution, of the 6.3 Tg N/

yr of U.S. NOx emissions, 2.7 Tg N/yr are deposited 

back onto the land and surface waters of the U.S. Thus, 

by difference we estimate that as much as 3.6 Tg N/

yr of the U.S. NOx emissions are advected out of the 

U.S. via the atmosphere. Similarly, of the 3.1 Tg N/yr 

of NH3 that are emitted into the U.S. atmosphere each 

year, about 2.1 Tg N/yr are deposited onto the land 

and surface waters of the U.S., and about 1 Tg N/yr is 

advected out of the U.S. via the atmosphere. Emissions 

of N2O discharge about 0.8 Tg N/yr into the global 

atmosphere. In sum, 5.4 Tg N are advected out of the 

U.S. from all sources each year either to other nations or 

to the global atmospheric or ocean commons.

Riverine discharges of Nr to the U.S. coastal zone 

account for 4.8 Tg N/yr, while export of N-containing 

commodities (e.g., grain) removes another 4.3 Tg N/

yr from the U.S.. Altogether, along with 5.4 Tg N/yr 

of atmospheric advection, these total Nr outputs out of 

the U.S. continental environment add up to about 14 

Tg N/yr, leaving about 21 Tg N/yr unaccounted for. Of 

this amount, we estimate that 5 Tg N/yr are stored in 

soils, vegetation, and groundwater and, by difference, 

we estimate that about 16 Tg N/yr are denitrified to N2. 

Denitrification, a process that microbially converts Nr 

to N2 (as well as forming some N2O) requires both a 

carbon source and anaerobic conditions, a situation that is 

found in wetlands, oxygen-depleted streams, rivers, and 

the hypolimnion of reservoirs (or their sediments), soils, 

and engineered denitrification systems. This process can 

be a major Nr sink in river basins. There are substantial 

uncertainties (+/- 50%) for estimated emission and 

deposition and terms that are arrived at by difference (e.g., 

atmospheric advection and denitrification) – especially 

those that involve NHx. The Committee considered 

these uncertainties in developing the “Overarching 

Recommendations” of this report.

Current EPA Nr risk management and 

research programs

The parts of EPA most directly concerned with 

managing or conducting research on Nr are the Office 

of Air and Radiation, the Office of Water, and the Office 

of Research and Development (ORD). Over a dozen 

programs of EPA’s Office of Air and Radiation reduce 

risks from Nr. These programs and related activities 

include: National Ambient Air Quality Standards 

standard setting and implementation; emission standards 

for industrial stationary sources and area sources; the 

Acid Rain Program; the Clean Air Interstate Rule; 

and programs that focus on mobile source emissions. 

Programs designed to save energy, such as Energy Star, 

tend to reduce emissions of Nr as well. EPA’s Office of 

Water addresses Nr under both the Clean Water Act and 

the Safe Drinking Water Act through activities such as: 

criteria development and standard setting; total maximum 

daily load (TMDL) development; National Pollution 

Discharge Elimination System (NPDE S) permits; 

infrastructure financing through the Drinking Water and 

Clean Water State Revolving Funds; watershed planning; 

wetlands preservation; and regulation of stormwater and 

runoff sources that include municipal separate storm 

sewer systems (MS4), and concentrated animal feeding 

operations (CAFOs). EPA’s Office of Research and 

Development aims to conduct leading-edge research and 

foster the sound use of science and technology in support 

of the Agency’s mission. The Office of Research and 

Development is well recognized for providing a scientific 

basis for the development of the National Ambient Air 

Quality Standards for NOx and particulate matter (PM). 

The Office of Research and Development’s Ecosystem 

Services Research Program has been developed to 

identify and quantify the positive and negative impacts 

on ecosystem services resulting from changes in nitrogen 

loadings from major source categories. This research 

will support policy and management decisions in EPA’s 

Offices of Air and Radiation and Water.

EPA has brought a great variety of risk reduction tools 

to bear on Nr: conventional regulation and enforcement; 

cap and trade approaches; measurement, monitoring and 

place-based approaches; control technology development 

and verification; communication and education; 

intergovernmental and international cooperation; and 

voluntary approaches. The variety and breadth of EPA 

programs addressing Nr reflect the ubiquity of Nr in the 

environment, the historical single-medium regulatory 

approach, and the lack of a “silver bullet” for reducing 

risks from Nr.

Need for an integrated management 

strategy

The EPA programs discussed above (and the programs 

of EPA’s predecessor organizations) have been active 

in the management of Nr through efforts to: decrease 

or transform Nr in sewage; control NOx to decrease 

photochemical smog and acid rain; control Nr inputs 

to coastal systems; control fine particulates in the 

atmosphere; and decrease Nr leaching and runoff from 

crop and animal production systems and developed 
ES-7

6 An integrated nitrogen management strategy takes a holistic approach for managing Nr. In the context of the nitrogen cascade, all Nr anthropogenic 

creation and destruction mechanisms and all Nr uses are recognized. The strategy should take account of synergies and trade-offs, to ensure that 

decreasing one problem related to nitrogen does not result in other unintended adverse environmental, economic and societal consequences. By 

identifying relative priorities, assessing cost effectiveness and risks, the strategy should seek to maximize the benefits of Nr, while limiting overall 

adverse effects.

lands. As beneficial as those efforts have been, they have 

focused on the specific problem without consideration of 

the interaction of a particular system with other systems 

downstream or downwind. Given the reality of the 

nitrogen cascade, this approach may result in short-term 

benefits for a particular system but may only temporarily 

delay larger-scale impacts on other systems. Thus there 

is a need to integrate N management programs, to ensure 

that efforts to lessen the problems caused by N in one area 

of the environment do not result in unintended problems 

in other areas.

Biofuels feedstock production provides a good 

example of the need for comprehensive and integrated 

assessment and management of Nr. Increasing corn 

production for ethanol has raised the prospect of increased 

Nr losses (i.e., transfer from fertilized land to water) and 

degraded water quality. The alternative of cellulosic based 

ethanol does not necessarily mitigate the potential for this 

negative externality. High yields of cellulosic materials 

also require N and the “marginal” land assumed for such 

production may be more susceptible to nutrient leakage. 

Another good example is provided in Chapter 4 of this 

report (Box 2). This example considers the water impacts 

of Nr in the Chesapeake Bay and shows that the total 

reduction of damage from excess Nr may rely nearly as 

much on stricter enforcement of the Clean Air Act as the 

Clean Water Act. This challenges the traditional approach 

to regulation, but it is a consequence of comprehensively 

examining Nr guided by the nitrogen cascade.

There can be many unintended consequences 

associated with a focus on managing one pollutant, even 

an integrated focus on various forms of N. For example, 

as further discussed in Chapter 4 and Appendix G of this 

report, numerous lakes, reservoirs, rivers, estuaries (e.g., 

the Gulf of Mexico), and fjords worldwide exhibit N and 

phosphorus (P) co-limitation, either simultaneously or 

in seasonally-shifting patterns. Therefore, strategies are 

needed to reduce both P and N inputs, and not all control 

practices will be effective for dual nutrient reduction. 

Synergistic effects on nutrient loss reductions can occur 

where combinations of control practices produce more 

or less than the sum of their individual reductions (U.S. 

EPA SAB, 2007). An integrated strategy should take this 

into consideration.

Objectives of the SAB Integrated 

Nitrogen Committee study

The EPA Science Advisory Board formed the 

Integrated Nitrogen Committee to assist EPA in its 

understanding and management of nitrogen-related 

air, land, and water pollution issues. In this report, the 

Committee has provided findings and recommendations 

addressing the following objectives. Assessment of 

the challenges and costs to EPA of implementing the 

recommendations is beyond the scope of the report.

1. Identify and analyze, from a scientific 

perspective, the problems Nr presents in the 

environment and the links among them.

To address this objective, the Committee used the 

nitrogen cascade framework to determine the major 

sources of newly created Nr in the U.S. (Figure ES-1). 

The flows of Nr within the food, fiber, feed and bioenergy 

production systems and developed lands in the U.S. were 

examined, paying special attention to the locations within 

each of these systems where Nr is lost to the environment. 

The same process was employed for fossil fuel energy 

production but, since all the Nr formed and released 

during energy production is lost to the environment, the 

Committee identified the important energy producing 

sectors that contribute to Nr emissions.

The Committee next examined the fate of the Nr 

lost to the environment, estimated the amount stored in 

different systems (e.g., forest soils) and tracked Nr as 

it is transferred from one environmental system (e.g., 

the atmosphere) to another (e.g., terrestrial and aquatic 

ecosystems).

Source and fate analyses set the stage for identifying 

the environmental and human health problems Nr presents, 

and the links among them. Using the nitrogen cascade, the 

Committee identified the impacts Nr has on people and 

ecosystem functions as it moves through each system. The 

Committee also addressed the alternative metrics that could 

be used, including the number of tons of specific forms of 

Nr, human health indicators and the economic damage cost, 

to assess incommensurable impacts due to environmental 

changes (e.g., acid deposition) vs. impacts due to losses of 

ecosystem services (e.g., loss of biodiversity), and tradeoffs among Nr impacts. 

2. Evaluate the contribution an integrated 

nitrogen management strategy6 could make to 

environmental protection.

An integrated management strategy should take 

into account the contributions of all Nr sources, and all 

chemical species of Nr that adversely impact both human 

health and environmental systems. Further, an integrated 

strategy should ensure that solving one problem related 

to Nr does not exacerbate another problem or diminish 

ecosystem services that support societal demands. In 

short, the strategy should seek to achieve desirable 

benefits of Nr, while limiting adverse effects.
ES-8

To address this challenge, the Committee identified 

several actions that could be taken to better manage 

Nr in one environmental system and avoid unintended 

consequences in another. Examples of “integrative” 

management actions that could be taken are highlighted in 

that discussion.

3. Identify additional risk management options 

for EPA’s consideration.

As further discussed below, the Committee has 

identified four major goals for management actions that 

collectively have the potential to decrease Nr losses 

to the environment by about 25 percent. Decreasing 

Nr emissions by these actions will result in further 

decreases in Nr-related impacts throughout the nitrogen 

cascade. The Committee has suggested several ways to 

attain these management goals including conservation 

measures, additional regulatory steps, voluntary actions, 

application of modern technologies, and end-of-pipe 

approaches. These are initial but significant actions; 

however, others should be taken once the recommended 

actions are completed and assessed, and further 

opportunities are explored in an adaptive management 

approach. Thus, the last sections of this report focus on a 

better understanding of Nr dynamics and impacts in the 

U.S. that could lead to more cost efficient management, 

balancing human and environmental needs.

4. Make recommendations to EPA concerning 

improvements in nitrogen research to support 

risk reduction.

In this report, the Committee has provided numerous 

recommendations for additional Nr research to support 

risk reduction activities. These research recommendations 

are discussed in various chapters of the report and 

are consolidated in the summary of findings and 

recommendations presented in Chapter 6.

Major Findings and Recommendations

Throughout the report there are boxes containing 

summary statements labeled “Findings.” Attached to these 

findings are one or more specific “Recommendations” for 

actions that could be taken by EPA or other management 

authorities. In each case, the intent is to provide the 

scientific foundation regarding a specific Nr-relevant 

environmental issue and one or more recommendations 

by which EPA acting alone or in cooperation with other 

organizations could use currently available technology to 

decrease the amount of Nr lost to the U.S. environment. 

The findings and recommendations are consolidated in 

Chapter 6 of this report.

Overarching recommendations 

Optimizing the benefits of Nr, and minimizing its 

impacts, will require an integrated nitrogen management 

strategy that involves action not only on the part of EPA, 

but also coordination with other federal agencies, the 

states, the private sector, universities, and the public, 

supported by a strong public outreach program. Therefore 

the Committee has also provided four overarching 

recommendations to assist EPA in its understanding and 

management of nitrogen-related air, land, and water 

pollution issues:

Overarching Recommendation 1: The Committee 

recommends an integrated approach to the management 

of Nr. This approach draws upon a combination of 

implementation mechanisms. Each mechanism must 

be appropriate to the nature of the problem at hand, 

be supported by critical research on decreasing the 

risks of excess Nr, and reflect an integrated policy that 

recognizes the complexities and tradeoffs associated with 

the nitrogen cascade. Management efforts at one point 

in the cascade may be more efficient and cost effective 

than control or intervention at another point. This is why 

understanding the nature and dynamics of the N cascade 

is critically important.

Overarching Recommendation 2: The framing 

of the reactive nitrogen cascade provides a means for 

tracking nitrogen as it changes form and passes through 

multiple ecosystems and media. This complexity 

requires the use of innovative management systems and 

regulatory structures to address the environmental and 

human health implications of the most damaging forms 

and quantities of Nr. It is difficult to create de novo 

fully effective regulations for such a complex system 

so we recommend utilizing adaptive management to 

continuously improve the effectiveness and lower the 

cost of implementation policies. This in turn will require 

a monitoring system that will provide feedback on the 

effectiveness of specific actions taken to lower fluxes and 

concentrations of Nr. 

Overarching Recommendation 3: An intraAgency Nr management task force within EPA is 

recommended to build on existing Nr research and 

management capabilities within the Agency. This 

task force should be aimed at increasing scientific 

understanding of: (1) Nr impacts on terrestrial and 

aquatic ecosystems, human health, and climate; (2) 

Nr-relevant monitoring requirements; and (3) the most 

efficient and cost-effective means by which to decrease 

various adverse impacts of Nr loads as they cascade 

through the environment. 

Overarching Recommendation 4: Successful Nr 

management will require changes in the way EPA 

interacts with other agencies. Coordinated federal 

programs could better address Nr concerns and 

help ensure clear responsibilities for monitoring, 

modeling, researching and managing Nr in the 

environment. Thus, the Committee recommends that 

EPA convene an inter-agency Nr management task 

force. It is recommended that the members of this 

inter-agency task force include at least the following 

federal agencies: U.S. Department of Agriculture, 
ES-9

U.S. Department of Energy, U.S. Department of 

Housing and Urban Development, U.S. Department 

of Transportation, National Oceanic and Atmospheric 

Administration, U.S. Geological Survey, U.S. Forest 

Service, and Federal Emergency Management Agency. 

The EPA Office of International and Tribal Affairs 

should work closely with the Department of State to 

ensure that EPA is aware of international efforts to 

control Nr and is developing national strategies that 

are compatible with international initiatives. Similar 

recommendations for coordination and joint action 

among and between agencies at both state and federal 

levels have been made in the National Research 

Council’s recent reports on the Mississippi Basin 

(NRC, 2008b, 2009). These intra- and inter-agency 

Nr management task forces should take a systems 

approach to research, monitoring, and evaluation to 

inform public policy related to Nr management, and 

implement a systems approach to Nr management, as 

recommended by the Committee.

Summary of specific recommendations 

by study objective 

The Committee’s findings and recommendations 

corresponding to each of the four study objectives are 

summarized briefly below. 

1. Identify and analyze, from a scientific 

perspective, the problems Nr presents in the 

environment and the links among them. 

The Committee finds that uncertainty associated 

with rapid expansion of biofuels, losses of Nr from 

grasslands, forests, and urban areas, and the rate and 

extent of denitrification have created the need to 

measure, model, and report all forms of Nr consistently 

and accurately. Addressing this need will decrease 

uncertainty in the understanding of the fate of Nr that 

is introduced into the environment and lead to a better 

understanding of the impacts of excess Nr on the health 

of people and ecosystems. This should be accomplished 

through a coordinated effort among cognizant federal 

and state agencies, and universities. 

In addition, the Committee recommends that EPA 

routinely and consistently account for the presence 

of Nr in the environment in forms appropriate to the 

medium in which they occur (air, land, and water) and 

that accounting documents be produced and published 

periodically (for example, in a fashion similar to 

National Atmospheric Deposition Program summary 

reports). The Committee understands that such an 

undertaking will require substantial resources, and 

encourages the Agency to develop and strengthen 

partnerships with appropriate federal and state 

agencies, and private sector organizations, with parallel 

interests in advancing the necessary underlying science 

of Nr creation, transport and transformation, impacts, 

and management. 

2. Evaluate the contribution an integrated 

nitrogen management strategy could make to 

environmental protection.

The Committee finds that effective management of Nr in 

the environment must recognize the existence of tradeoffs 

across a number of impact categories involving the cycling 

of nitrogen and other elements. In addition, an integrated 

multi-media approach to monitoring Nr is needed. 

In that regard, the Committee recommends that: 

1. EPA should develop a uniform assessment and 

management framework that considers the effects of 

Nr loading over a range of scales reflecting ecosystem, 

watershed, and regional levels. The framework should 

include all inputs related to atmospheric and riverine 

delivery of Nr to estuaries, their comprehensive effects 

on marine eutrophication dynamics, and their potential 

for management. 

2. EPA should examine the full range of traditional and 

ecosystem response categories, including economic and 

ecosystem services, as a basis for expressing Nr impacts 

in the environment, and for building better understanding 

and support for integrated management efforts. 

3. Identify additional risk management options 

for EPA’s consideration.

The Committee finds that a number of risk 

management actions should be considered to reduce Nr 

loading and transfer to the environment. These include 

farm-level improvements in manure management, actions 

to reduce atmospheric emissions of Nr, and interventions 

to control Nr in water management programs. As an 

example, the Committee recommends that EPA should 

reexamine the criteria pollutant “oxides of nitrogen” and 

the indicator species NO2 and consider supplementing this 

with NHx and NOy as indicators of chemically reactive 

nitrogen (Nr without N2O). 

4. Make recommendations to EPA concerning 

improvements in nitrogen research to support 

risk reduction.

The Committee finds that research is needed in a 

number of areas to support Nr risk reduction activities. 

These areas include research to advance the understanding 

of: the quantity and fate of Nr applied to major crops; 

how to accelerate crop yields while increasing N fertilizer 

uptake efficiency; agricultural emissions of forms of 

Nr; atmospheric deposition of Nr; and the potential for 

amplification of Nr-related climate impacts. 

Four recommended management 

actions 

Consistent with the overarching and specific 

recommendations noted above, the Committee identified 

four management actions that could be undertaken in 

the near term by applying existing proven science and 

technology and determined how those actions could 

contribute to the reduction of excess Nr in the environment. 
ES-10

1. The Committee estimates that if EPA were to expand 

its NOx control efforts for emissions of mobile sources 

and power plants, a 2.0 Tg N/yr decrease in the 

generation of reactive nitrogen could be achieved. Such 

changes can be effected by applying existing, proven 

technology. Emissions from many point sources are 

controlled with low-NOx burners or NOx reduction.

Such equipment should also be installed on industrial 

boilers and the remaining, uncontrolled power plants. 

NOx controls for modern on-road vehicles are effective 

and these technologies should be applied to off-road 

vehicles, locomotives, ships and other devices with 

internal combustion engines.

2. The Committee estimates that excess flows of Nr into 

streams, rivers, and coastal systems can be decreased 

by approximately 20% (approximately 1 Tg N/yr) 

through improved landscape management and without 

undue disruption to agricultural production. This would 

include activities such as using large-scale wetland 

creation and restoration to provide needed ecosystem 

services of Nr retention and conversion as well as 

matching cropping systems and intensity of Nr use to 

land characteristics. Improved tile-drainage systems 

and riparian buffers on cropland, and implementing 

stormwater and non-point source management practices 

(e.g., EPA permitting and funding programs) are 

important components. In addition, the Committee 

estimates that crop N-uptake efficiencies can be 

increased by up to 25% over current practices through 

a combination of knowledge-based practices and 

advances in fertilizer technology (such as controlled 

release and inhibition of nitrification). Crop output can 

be increased while decreasing total Nr by up to 20% of 

applied artificial Nr, amounting to ~2.4 Tg N/yr below 

current amounts of Nr additions to the environment. 

These are appropriate actions that could be taken with 

today’s available technologies and further progress is 

possible.

3. The Committee estimates that livestock-derived NH3

emissions can be decreased by 30% (a decrease of 0.5 

Tg N/yr) by a combination of BMPs and engineered 

solutions. This is expected to decrease PM2.5 by 

approximately 0.3 micrograms per cubic meter (2.5%), 

and improve health of ecosystems by achieving 

progress towards critical load recommendations. 

Additionally we estimate that NH3 emissions derived 

from fertilizer applications can be decreased by 

20% (decrease by approximately 0.2 Tg N/yr), 

through BMPs that focus on improvements related to 

application rate, timing, and placement.

4. The Committee recommends that a high priority be 

assigned to increasing funding for nutrient management. 

We estimate that adequate financial support for sewage 

treatment infrastructure upgrades to remove nutrients 

could decrease Nr emissions by between 0.5 and 0.8 

Tg N/yr. Additional Nr management from eligible 

stormwater and nonpoint sources could be accomplished 

through increased support.

Implementing these suggestions will decrease the 

amount of Nr introduced into the United States by about 

25%, which will similarly decrease the amount of Nr 

lost to the atmosphere, soils and waters. The Committee 

believes that these represent realistic and attainable 

near-term outcomes, however further reductions are 

undoubtedly needed for many N-sensitive ecosystems and 

to ensure that health-related standards are maintained. 
1

SABReactive Nitrogen 

in the United States 

An Analysis of Inputs, Flows, Consequences, 

and Management Options
2
3

1

Introduction

1.1. Overview of the Problem – Impacts 

of Excess Nr on Human Health and the 

Environment

Nitrogen is an essential nutrient that governs the 

growth and reproduction of living organisms. Reactive 

nitrogen (Nr), in contrast to non-reactive gaseous N2, 

includes all biologically active, chemically reactive, 

and radiatively active nitrogen (N) compounds in the 

atmosphere and biosphere of the Earth. Anthropogenic 

creation of Nr provides essential benefits for humans – 

first and foremost in meeting human dietary needs. In fact, 

a large proportion of the human population of the earth 

could not be sustained if synthetic nitrogen fertilizers did 

not augment food production significantly all over the 

world. However, excess releases of Nr to the environment 

from human activities such as fossil fuel combustion and 

agriculture are a major cause of air and water quality 

degradation that has been linked to significant impacts on 

human and ecosystem health. 

Reactive nitrogen (Nr) includes inorganic chemically 

reduced forms of N (NHx) [e.g., ammonia (NH3) and 

ammonium ion (NH4+)], inorganic chemically oxidized 

forms of N [e.g., nitrogen oxides (NOx), nitric acid 

(HNO3), nitrous oxide (N2O), N2O5, HONO, peroxy 

acetyl compounds such as peroxyacytyl nitrate (PAN), 

and nitrate ion (NO3-)], as well as organic compounds 

(e.g., urea, amines, amino acids, and proteins).

The negative consequences of Nr flux in the U.S. 

environment include increases in photochemical smog 

and atmospheric particulate matter (PM2.5), decreases 

in atmospheric visibility, both increases and decrease 

in productivity of grasslands and forests, acidification 

of soils and freshwaters, accelerating estuarine and 

coastal eutrophication, increases in the emission of 

greenhouse gases (GHG) to the atmosphere, and decreases 

in stratospheric ozone concentrations. Most of these 

changes in environmental conditions lead to a variety of 

negative impacts on both ecosystem and human health 

(Johnson and Siccama, 1983; Heck et al., 1984; Paerl, 

1988; MacKenzie and El-Ashry, 1990; ECOH AB, 1995; 

Bricker et al., 1999; Rabalais et al., 1999; NRC, 2000; 

Mitsch et al., 2001; Fenn et al., 2003; Ezzati et al., 

2004; Mokad et al., 2004; Verity et al., 2006; U.S. EPA 

Clean Air Scientific Advisory Committee, 2008; U.S. 

EPA SAB, 2008; Bobbink et al., 2010). In light of the 

magnitude of the human alteration of the nitrogen cycle, 

and the resulting negative consequences on humans 

and ecosystems, the National Academy of Engineering 

has identified management of the nitrogen as one of the 

“grand challenges” facing this country (National Academy 

of Engineering, 2008).

1.2. The Nitrogen Cascade – Nr Loading, 

Cycling, and Exposure

Approximately 78% of the atmosphere is diatomic 

nitrogen (N2), which is unavailable to most organisms 

because of the strength of the triple bond that holds the two 

N atoms together. Over evolutionary history, only a limited 

number of species of bacteria and archaea have evolved 

the ability to convert N2 to Nr via biological N fixation. 

Thus, even with adaptations to use N efficiently, many 

ecosystems of the world are limited by N.

Anthropogenic creation of Nr

Nitrogen limitation of ecosystem production has 

driven humans to use increasingly sophisticated and 

energy-intensive measures to obtain Nr to sustain food 

production and to produce other commodities (e.g., 

nylon, explosives). In pre-history, hunters and gatherers 

harvested food from natural stocks. With the advent of 

agriculture, local sources of Nr were used (soil stocks, 

crop residue, and manures) to increase productivity 

of landscapes. In the nineteenth century, long-range 

transport of Nr to sustain food production increased 

with the shipment of bird guano from the Pacific Islands 

and nitrates from South America to Europe and other 

locations. By the beginning of the twentieth century, these 

sources were not sufficient to sustain the growing global 

population requirements for food. 

This deficiency led to what has been called one of 

the world’s most important discoveries – how to extract 

N2 from the atmosphere and convert it to ammonia 

(NH3) – called the Haber-Bosch process (Smil, 2001; 

Erisman et al., 2008). Today, this process and cultivationinduced biological N fixation (C-BNF) introduce over 

140 teragrams (Tg) of N per year (hereafter expressed 

as Tg N/yr) into the global environment to increase food 

production (Galloway et al., 2008). Another 23 Tg N/

yr are introduced by the Haber-Bosch process for the 

chemical industry, and 25 Tg N/yr are introduced via the 

combustion of fossil fuels (Galloway et al., 2008). 

The total global anthropogenic Nr creation rate is ~190 

Tg N/yr (2005), substantially larger than the median of 

estimates for Nr creation by natural terrestrial processes 

(~100 Tg N/yr) (Galloway et al., 2008). The fact that 

humans are more effective than nature in Nr creation 

means that on average, humans are less reliant on natural 

sources of Nr. However, with global commodity stocks 

running at a 58-day supply and food prices increasing 
4

dramatically, the challenge is to increase the nutrient use 

efficiency of Nr in agricultural systems while maintaining 

or increasing yields (USDA, ERS/World Agricultural 

Outlook Board, July 11, 2008. World Agricultural Supply 

and Demand Estimates).

There are large regional disparities in Nr creation rates 

on both absolute and per capita bases. Total Nr creation 

is larger in Asia than in any other region. Per capita Nr 

creation is largest in North America and Europe. Humans 

also redistribute large amounts of Nr among countries 

or regions of the world through exports of fertilizers, 

feed grains, and fossil fuels. Nevertheless, there are large 

regions of the world with populations approaching one 

billion, where there is malnutrition in part due to a lack of 

available Nr to sustain crop production.

The introduction of Nr into most regions of the United 

States by humans has greatly increased food availability. 

However, since essentially all the Nr created for food 

production and by fossil fuel combustion is lost to the 

environment, it has also greatly increased the contribution 

of Nr to a wide variety of environmental problems. 

Most plants, animals, and microorganisms are adapted 

to efficiently use and retain Nr. Addition of Nr to most 

ecosystems may first lead to increased uptake, growth, 

and storage – and hence to increased biomass, including 

food or fiber production. However, further addition of 

Nr in excessive amounts often leads to imbalances in the 

movement of Nr among reservoirs and potential losses7

to the environment in the form of air emission or water 

discharges into other ecosystems where Nr may disrupt 

ecosystem functions and have a negative impact on 

resources. In essence, the assimilative capacity of the 

ecosystem may be insufficient to benefit from increases in 

Nr without disruptive changes.

These changes, which impact air, land, water and the 

balance of life in an interrelated fashion, are often referred 

to as a cascade of effects from excess Nr8 or the “nitrogen 

cascade” (Figure 1). Unlike other element-based pollution 

problems, the N cascade links the negative impacts, where 

one N-containing molecule can in sequence contribute to 

all the environmental issues mentioned above.

The nitrogen cascade has three dimensions: 

biogeochemical, alterations in the environment, and 

human and ecosystem consequences. 

The “biogeochemical” dimension of the nitrogen 

cascade involves: Nr creation from N2 as a consequence 

of chemical, food and energy production; Nr use in food 

and chemical production; Nr losses to the environment; 

changes in Nr species residence times in environmental 

reservoirs; Nr transfers among reservoirs; and Nr 

conversion back to N2. Alterations to the environment 

then result from increased Nr levels in the environment. 

These alterations have negative consequences for 

ecosystem and human health at local, regional, national 

and global scales. Because nitrogen is a critical resource 

and also a contributor to many of the environmental 

concerns facing the U.S. today, it is imperative to 

understand how human action has altered N cycling in the 

U.S., and the consequences of those alterations on people 

and ecosystems. The overarching question is, how do 

we protect and sustain ecosystems that provide multiple 

benefits to society while also providing the interconnected 

material, food and energy required by society?

Nr inputs to the nation and the world have been 

increasing, largely due to human activities associated 

with food production and fossil fuel combustion. Despite 

the obvious benefits of a plentiful supply of food and 

energy, the adverse consequences associated with the 

accumulation of Nr in the environment are large, with 

implications for human health and the environment. 

The greater the inputs of Nr to the landscape, the 

greater the potential for negative effects, caused by 

greenhouse gas (GHG) production, ground level ozone, 

acid deposition, and Nr overload that can contribute 

to climate change, degradation of soils and vegetation, 

acidification of streams, lakes and rivers, estuarine and 

coastal eutrophication, hypoxia and habitat loss.

The growing nature of the Nr problem, and the adverse 

and intertwined consequences associated with Nr inputs 

to air, land, and water as exhibited in the N cascade 

underscore the need for researchers and managers to 

explore integrated strategies that minimize N inputs, 

maximize its use efficiency, promote Nr removal 

processes and protect humans and natural resources. 

The concept of the nitrogen cascade highlights that 

once a new Nr molecule is created, it can, in sequence, 

travel throughout the environment contributing to major 

environmental problems (Galloway et al., 2003). The 

adaptation of the cascade in Figure 1 was developed by 

the SAB Integrated Nitrogen Committee (INC) to provide 

a context for considering nitrogen-related issues and 

ecosystem effects in the U.S. To consider the cascading 

effects of Nr in the U.S., we examined the various 

atmospheric, terrestrial, and aquatic environmental 

7 In the context of this report, “losses” refers to transfers among systems and not the conversion of Nr to N2. Whenever N2 formation is discussed, it is 

explicitly stated.

8 Excess reactive nitrogen (Nr) is defined as the amount of Nr that is present in, or introduced into, an environmental system (e.g., Nr inputs to the 

atmosphere, Nr inputs to grasslands and forests, Nr inputs to estuaries) from anthropogenic sources that is not incorporated into agricultural and 

other biological products (e.g., food, feed, fuel, and fiber), or stored in long-term storage pools (e.g., cropland soils).Thresholds are used to determine 

the amount of excess Nr that causes negative effects on ecosystem services and functions, and human health. Thresholds vary by metric (e.g., 

concentration, loading, etc) and depend on the environmental system (e.g., atmosphere, forest). Examples for specific thresholds are given later in the 

report in relevant sections.
5

systems where Nr is stored, and the magnitudes of 

the various flows of N to, from, and within them. The 

nitrogen cascade concept implies the cycling of Nr among 

these systems. The process of denitrification is the only 

mechanism by which Nr is converted to chemically 

inert N2, “closing” the continuous cycle (Figure 1 shows 

only flows of reactive nitrogen, not N2). Denitrification 

can occur in any of the indicated reservoirs except the 

atmosphere.

The “new” N box in Figure 1 depicts the two primary 

anthropogenic sources by which Nr originates – energy 

production and food production – and where Nr from 

these sources enter ecosystems. Energy production 

includes both fossil fuel and biofuel combustion. Food 

production includes N fertilizer produced in the U.S., 

cultivation-induced biological N (C-BNF) in the U.S., 

production of animals and crops in the U.S. for human 

consumption, and imports of N-containing fertilizer, grain 

and meat to the U.S. 

The atmospheric system box in Figure 1 indicates that 

tropospheric concentrations of both ozone and particulate 

matter are increased due to emissions of nitrogen oxides9

(NOx) to the atmosphere. The ovals illustrate that the 

increase in N2O concentrations, in turn, contribute to 

the greenhouse effect in the troposphere and to ozone 

depletion in the stratosphere. Except for N2O, there is 

limited Nr storage in the atmosphere. Losses of Nr from 

the atmospheric system include total oxidized nitrogen10

(NOy), reduced nitrogen11 (NHx), and organic nitrogen 

(Norg) deposition to terrestrial and aquatic ecosystems of 

the earth’s surface. There is little potential for conversion 

of Nr to N2 via denitrification in air. However, once 

airborne deposition of Nr occurs it will be subject to 

denitrification pathways via soil and water.

The terrestrial system box in Figure 1 depicts that 

Nr enters agricultural lands via food production and 

is introduced to the entire terrestrial landscape via 

atmospheric deposition. Within agricultural regions there 

is cycling among soils, crops and animals, and then a 

transfer of Nr as food to populated regions, from which 

there are Nr losses to the environment (e.g., sewage, 

Figure 1: The nitrogen cascade

The Nitrogen Cascade Atmospheric

Terrestrial

Aquatic

Energy Production

& combustion of fossil fuels

New Nitrogen

Food Production

& creation of synthetic fertilizers

NOy, NHx, Norg

NOy, NHx, Norg

stratosphere

ozone troposphere

ozone depletion

greenhouse gases

particulate matter

crops animals

soils

people

soils

landscape

forests grasslands

soils

Agricultural Populated

Vegetated

Groundwater

Oceans

Coastal Bays

& Estuaries

Surface Water

& Wetlands

NOy, NHx, Norg

N2O

N2O

NOx

NH3

Norg

biogeochemical cycling

ecosystem productivity

acidification

eutrophication

NOx

NHx

9 NOx (oxides of nitrogen) includes NO + NO2

10 NOy (total oxidized nitrogen) includes NO, NO2, NO3, N2O5, HONO, HNO3, NO3

-

, PAN and other organo-nitrates, RONO2

11 NHx (reduced nitrogen) includes NH3 + NH4
6

landfills). The ovals showing ecosystem productivity 

and biogeochemical cycling reflect that Nr is actively 

transported and transformed within the terrestrial system, 

and that as a consequence there are significant impacts 

on ecosystem productivity due to fertilization and 

acidification, often with resulting losses of biodiversity. 

There is ample opportunity for Nr storage in both biomass 

and soils. Losses of Nr from this system occur by leaching 

and runoff of NOy, NHx and Norg to aquatic ecosystems 

and by emissions to the atmospheric system as NOx, NH3, 

Norg, and N2O. There is potential for conversion of Nr to 

N2 via denitrification in the terrestrial system. 

The aquatic system box in Figure 1 shows that Nr 

is introduced via leaching and runoff from terrestrial 

ecosystems and via deposition from atmospheric 

ecosystems. Connected with the hydrological cycle, there 

are Nr fluxes downstream with ultimate transport to coastal 

systems. Within the aquatic system, the ovals highlight 

two significant impacts of waterborne Nr acidification of 

freshwaters and eutrophication of fresh and coastal waters. 

Except for Nr accumulation in groundwater reservoirs, 

there is limited Nr storage within the hydrosphere. Losses 

of Nr from the aquatic system are primarily via N2O

emissions to the atmospheric system. There is a very large 

potential for conversion of Nr to N2 via denitrification in 

water and wetlands. 

NOy, NHx and N2O are all components of Nr, but 

a fundamental difference is that the NOy and NHx are 

rapidly transferred from the atmosphere to receiving 

ecosystems due to a short atmospheric residence time 

(≤ 10 days) where they continue to contribute to the N

cascade (Galloway et al., 2004). Because of its longer 

residence time (~100 years) however, N2O remains in the 

troposphere where it contributes to climate change, until 

it is transferred to the stratosphere, where it contributes to 

ozone depletion (Galloway et al., 2004).

1.3. EPA Activities to Manage Risks 

Posed by Nr

EPA activities to manage the risks posed by reactive 

nitrogen can be linked to the Agency’s broad strategic 

goals. EPA’s mission is to protect human health and 

the environment. In achieving this mission, EPA is 

accountable for addressing five goals given in the 2006 – 

2011 EPA Strategic Plan (U.S. EPA, 2006d): 

1. Clean air and global climate change 

2. Clean and safe water 

3. Land preservation and restoration

4. Healthy communities and ecosystems

5. Compliance and environmental stewardship

The Strategic Plan includes targets for reducing risk 

from N. EPA’s Report on the Environment (U.S. EPA, 

2008c), provides “data on environmental trends,” to 

determine whether or not EPA is on track to meet its 

targets and goals. EPA is responsible and accountable for 

reducing at least some risks from Nr.

As previously discussed, the principal mechanisms 

for Nr removal from circulation in the environment are 

complete denitrification (re-conversion of Nr back to nonreactive gaseous N2), and storage in long-term reservoirs 

(e.g., soils, sediments, and woody biomass). In some cases, 

it may be possible to capture Nr emissions or discharges 

and deliver them to food or fiber production areas where 

there are nitrogen deficiencies. However, as previously 

noted, major challenges in the management of the N cycle 

are how to decrease creation of Nr while still meeting 

societal needs, promote denitrification of excess Nr 

(without producing N2O), and improve the efficiency of use 

and reuse of excess Nr in a cost-effective manner. Solving 

these challenges will result in less Nr accumulation.

The parts of EPA most directly concerned with 

managing or conducting research on Nr are the Office of 

Air and Radiation, the Office of Water, and the Office of 

Research and Development. Programs designed to save 

energy, such as Energy Star, tend to reduce emissions of 

Nr as well. In over a dozen programs, EPA’s Office of Air 

and Radiation reduces risks from Nr. These programs and 

related activities include: 

n National Ambient Air Quality Standards (NAAQS) 

standard setting and implementation;

n Emission standards for industrial stationary sources and 

area sources

n Acid Rain Program

n Clean Air Interstate Rule

n Programs that focus on mobile source emissions 

EPA’s Office of Water addresses Nr under both the 

Clean Water Act and the Safe Drinking Water Act through 

activities such as: 

n Criteria development and standard setting

n Total maximum daily load (TMDL) development

n National Pollution Discharge Elimination System 

(NPDES) permits

n Watershed planning

n Wetlands preservation

n Regulation of concentrated animal feeding operations 

(CAFOs). 

EPA’s Office of Research and Development aims to 

conduct leading-edge research and foster the sound use of 

science and technology in support of EPA’s mission. The 

Office of Research and Development is well recognized 

for providing a scientific basis for the development of 

the National Ambient Air Quality Standards for NOx

and particulate matter (PM). The Office of Research and 

Development’s Ecosystem Services Research Program 

has been developed to identify and quantify the positive 
7

and negative impacts on ecosystem services resulting 

from changes in nitrogen loadings from major source 

categories to support policy and management decisions in 

EPA’s Offices of Air and Radiation and Water.

EPA has brought a great variety of risk reduction tools 

to bear on reactive N: 

n Conventional regulation and enforcement

n Cap and trade approaches

n Measurement, monitoring and place-based approaches

n Control technology development and verification

n Communication and education

n Intergovernmental and international cooperation

n Voluntary approaches

The variety and breadth of EPA programs addressing 

Nr reflect the ubiquity of Nr in the environment, the 

historical single medium regulatory approach, and the 

lack of a silver bullet for reducing risks from Nr. 

Need for an Integrated Nitrogen 

Management Strategy

The EPA programs discussed above (and the programs 

of EPA’s predecessor organizations) have been active in 

the management of Nr through efforts to: decrease the Nr 

amount in sewage, control NOx to decrease photochemical 

smog and acid rain, control Nr inputs to coastal 

systems, control fine particulates in the atmosphere, and 

decrease Nr leaching and runoff from crop and animal 

production systems. As beneficial as those efforts have 

been, they have focused on the specific problem without 

consideration of the interaction of their particular system 

with other systems downstream or downwind. Given the 

reality of the nitrogen cascade, this approach may result 

in short-term benefits for a particular system but will also 

likely only temporarily delay larger-scale impacts on other 

systems. Thus, there is a need to integrate N management 

programs, to ensure that efforts to lessen the problems 

caused by N in one area of the environment do not result 

in unintended problems in other areas. 

Biofuels feedstock production is a good example 

of this. Increasing corn production for ethanol raised 

the prospect of increased Nr losses and degraded water 

quality. The alternative of cellulosic based ethanol does 

not necessarily mitigate the potential for this negative 

externality. High yields of cellulosic materials also require 

N and the “marginal” land assumed for such production 

may be more susceptible to nutrient leakage (NRC, 2008a). 

In addition, there can be unintended consequences 

associated with a focus on one pollutant, even an integrated 

focus on various forms of nitrogen. For example, as further 

discussed in Chapter 4 and Appendix G of this report, 

numerous lakes, reservoirs, rivers, estuaries (e.g., the Gulf 

of Mexico), and fjords worldwide exhibit N and P colimitation, either simultaneously or in seasonally-shifting 

patterns. Therefore, strategies are needed to reduce both P

and N inputs, and not all control practices will be effective 

for dual nutrient reduction. There can be synergistic 

effects on nutrient loss reductions where combinations of 

control practices can produce more or less than the sum of 

their individual reductions (U.S. EPA SAB, 2007) and an 

integrated strategy should take this into consideration.

1.4. SAB Integrated Nitrogen Committee 

Study Objectives

The EPA Science Advisory Board has previously 

provided advice concerning management of nitrogenous 

compounds as well as integrated environmental decision 

making. In 1973, the Science Advisory Board issued 

a report Nitrogenous Compounds in the Environment

(U.S. EPA SAB, 1973). The report addressed sources and 

effects of nitrogenous compounds, including those from 

air emissions, animal wastes, crop agriculture, industrial 

processes, and solid wastes. The SAB concluded that, 

“At present, all known trends appear to be ones that 

can be managed and kept within control, if appropriate 

steps are taken now,” and provided recommendations 

relating to Nr research and control. In its 2000 report, 

Toward Integrated Environmental Decision-Making (U.S. 

EPA SAB, 2000) the SAB articulated a framework for 

integrated environmental decision-making. In that report, 

the SAB noted that the three-phased structure of the 

framework (problem formulation, analysis and decisionmaking, followed by implementation and evaluation) 

“belies the complexities involved in putting the concept 

of integrated decision-making into practice.” The SAB’s 

interests in N science and integrated environmental 

protection converged in 2007, when the SAB identified 

integrated N research and control strategies as an 

important issue facing the Agency and formed the 

Integrated Nitrogen Committee (the Committee) to 

conduct this study.

The Committee was charged by the Science Advisory 

Board to address the following four objectives:

1. Identify and analyze, from a scientific perspective, the 

problems reactive nitrogen presents in the environment 

and the links among them; 

2. Evaluate the contribution an integrated nitrogen 

management strategy12 could make to environmental 

protection;

12 An integrated nitrogen management strategy takes a holistic approach for managing Nr. In the context of the nitrogen cascade, all Nr anthropogenic 

creation and destruction mechanisms and all Nr uses are recognized. The strategy should take account of synergies and trade-offs, to ensure that 

decreasing one problem related to nitrogen does not result in other unintended adverse environmental, economic and societal consequences. By 

identifying relative priorities, assessing cost-effectiveness and risks, the strategy should seek to maximize the benefits of Nr, while limiting overall 

adverse effects.
8

3. Identify additional risk management options for EPA’s 

consideration; and

4. Make recommendations to EPA concerning 

improvements in nitrogen research to support 

risk reduction.

In this report the Committee has provided findings 

and recommendations addressing the study objectives. 

We recognize that there will be challenges and costs 

associated with the pursuit of the recommended 

management strategies, and that EPA requires 

statutory authority to take regulatory action. However, 

assessment of the challenges and costs to the Agency of 

implementing the recommendations is beyond the scope 

of this report.

1.5. Study Approach and Structure of 

the Report

To address the four objectives of this study, the 

Committee completed the following activities:

1. The Committee used the nitrogen cascade framework 

to determine the major sources of newly created Nr 

in the U.S. The flows of Nr within the food, fiber, 

feed, and bioenergy production systems of the U.S. 

were examined, paying special attention to the 

locations in each of these systems where Nr is lost 

to the environment. The same process was employed 

for energy production but, since all of the Nr formed 

during energy production is lost to the environment, the 

Committee identified the important energy producing 

sectors that contribute to Nr formation.

2. The Committee examined the fate of Nr lost to the 

environment, estimated the amount stored in different 

systems (e.g., forest soils), and tracked Nr as it is 

transferred from one environmental system (e.g., the 

atmosphere) to another (e.g., terrestrial and aquatic 

ecosystems).

3. Using the nitrogen cascade, the Committee identified 

the impacts Nr has on people and ecosystem functions 

as it moves through different systems. 

4. The Committee identified actions that could be 

taken based on available science and management 

practices to improve the integrative management of 

N. The Committee suggested ways in which each of 

these actions could be accomplished and estimated 

that together they could decrease Nr losses to the 

environment by about 25%. 

5. The Committee identified research needed to improve 

the scientific foundation to support specific Nr risk 

reduction activities.

Four public meetings were held during the course 

of the study and briefings were presented to the 

Committee by: EPA’s Office of Air and Radiation, Office 

of International Affairs, and Office of Water; the U.S. 

Department of Agriculture’s Agricultural Research 

Service, Cooperative State Research, Extension and 

Education Service, and the Economic Research Service; 

and external organizations such as the Energy Research 

Centre of the Netherlands, Environmental Defense 

Fund, International Plant Nutrition Institute, Iowa 

State University, LiveFuels, and the Soil and Water 

Conservation Society.

Additionally, the Committee invited scientists and 

managers from EPA, other federal agencies, states and 

localities, academia, non-governmental organizations, and 

the private sector to participate in an October 20-22, 2008 

workshop and meeting on Nitrogen Risk Management 

Integration. The purpose of the workshop was to receive 

public input on several subjects: the Committee’s 

preliminary assessment of Nr problems, consequences, 

and remedies, with emphasis on risk reduction; the 

Committee’s quantitative estimates of attainable Nr 

reductions; and mechanisms whereby the Nr strategy 

might be enacted. The Committee took this public input 

into consideration as it developed this report.

Structure of the report

This report contains six chapters. The report was 

developed for a multifaceted audience of scientists and 

policy makers and therefore the level of detail varies 

in different sections of the document. The introductory 

chapter provides an overview of problems caused 

by excess reactive nitrogen and describes the study 

objectives and approach. Chapters 2-6 discuss how the 

Committee has addressed the four study objectives and 

present specific findings and recommendations. The 

findings and recommendations corresponding to each of 

the study objectives are consolidated in Chapter 6.

n Study objective 1 (identification and analysis of the 

problems nitrogen presents in the environment and 

linkages among these problems) is addressed in 

Chapters 2 and 3. Chapter 2 focuses on the sources, 

transfer, and transformation of reactive nitrogen in 

environmental systems. Chapter 3 describes the impacts 

of reactive nitrogen on aquatic, atmospheric, and 

terrestrial ecosystems.

n Study objective 2 (evaluation of the contribution an 

integrated nitrogen management strategy could make 

to environmental protection) is addressed in Chapters 

4 and 5. Chapter 4 reviews the implications for risk 

reduction strategies for reactive nitrogen. Chapter 5 

discusses integrated risk reduction strategies.

n Study objective 3 (identification of additional risk 

management options for EPA’s consideration) is 

addressed in Chapters 5 and 6. In Chapter 6, the 

Committee identifies specific management goals 

for reducing the loss of reactive nitrogen to the 

environment. The Committee believes that these 

represent realistic near-term management goals that 

can be attained using current technology. However, the 

Committee emphasizes that further reduction beyond 
9

these goals will be needed to protect many N-sensitive 

ecosystems and to ensure that health-related standards 

are maintained.

 The Committee finds that the management goals could 

be attained by conservation measures, additional 

regulation, and application of modern technologies.

n Study objective 4 (recommendation of improvements 

in reactive nitrogen research to support risk 

reduction) is addressed in all of the report chapters 

and Chapter 6 contains a section describing the need 

for a comprehensive program to monitor Nr in the 

environment. 

Throughout this report there are boxes 

containing summary statements labeled “Findings.” 

Attached to these findings are one or more specific 

“Recommendations” for actions that could be taken by 

EPA or other management authorities. 
10
11

2

Sources, Transfer, and Transformation of 

Nr in Environmental Systems

The Committee was charged with identifying and 

analyzing, from a scientific perspective, the problems Nr 

presents in the environment and the links among them. 

This chapter addresses two aspects of the Committee’s 

work. The first aspect is the introduction of Nr into U.S. 

environmental systems from fossil fuel combustion 

and from food production, and the second aspect is the 

fate of Nr after it is emitted to the atmosphere by fossil 

fuel combustion or lost to the air, water and soils from 

agricultural production systems. The Nr budgets and 

calculations for the U.S. exclude Alaska and Hawaii 

because data were not available for these areas. Most of 

the Nr introduced into the environment comes from the 

Haber-Boesch process. Haber-Boesch contributes three to 

four times the Nr introduced from fossil fuel combustion 

and most of the Haber-Boesch introduction of Nr comes 

through agriculture.

2.1. Nr Flux in the Environment

As stated previously, although N is a major required 

nutrient that governs growth and reproduction of living 

organisms, Nr losses to the environment from human 

sources have a profound effect on air, water and soil 

quality. Human consumption of energy to sustain 

economic development results in emissions of NOx to 

the atmosphere via fossil fuel combustion. Consumption 

of food to meet nutritional requirements of a growing 

population results in agricultural emissions of NH3, urban 

and industrial emissions of NOx and N2O, and losses 

of NO3

-

 and other N compounds to water bodies due to 

leaching and runoff. Once released into the atmosphere by 

either human or natural processes, these Nr compounds 

undergo transformation through atmospheric reactions 

(e.g., gas-to-particle conversion), transport associated 

with wind, and finally wet and dry deposition. Reactive 

nitrogen lost from agricultural and peopled systems can 

enter groundwater, streams, lakes, estuaries, and coastal 

waters where the Nr can also undergo transformation 

mediated by a wide range of biotic and abiotic processes. 

The introduction of Nr into agroecosystems provides much 

of the world’s food. The losses of Nr to the environment 

throughout the food production process and during 

fossil fuel combustion contribute to many of the major 

environmental problems of today. The impacts of Nr on 

humans and ecosystems are discussed in the Executive 

Summary and Chapter 3 of this report.

Some key issues concerning management of Nr in 

the U.S. environment are the introduction of new Nr 

by imports, fertilizer production, cultivation-induced 

biological nitrogen fixation (C-BNF), and fossil fuel 

combustion. Other important issues are the distribution of 

Nr within agricultural systems and populated systems and 

redistribution of Nr through losses from those systems to 

the environment (Figure 1). National-level values for Nr 

fluxes are displayed in Table 1. Fluxes that represent the 

introduction of new Nr into the U.S. are marked with an 

asterisk and illustrated in Figure 1. In specific sections of 

this report these values have been used to more clearly 

determine the flux and fate of Nr in the U.S. Figure 2 

Illustrates sources of Nr introduced into the United States.

Figure 2: Sources of reactive nitrogen (Nr) introduced into the United States in 2002 (Tg N/yr).

4.2 6.4

3.8

1.9

7.7

10.9

Transportation

Stationary

Cultivation BNF

Haber Bosch N Fertilizer

Nonfertilizer Haber Bosch N 

BNF

Agriculture

Natural

Fossil Fuel

Industry
12

*Terms with an asterisk indicate Nr that is created, highlighting where reactive nitrogen is introduced to the environment.

Table 1: Nr fluxes for the United States, Tg N in 2002.a

Nr inputs to the Atmospheric environmental system Tg N/yr* %

 N2O-N emissions1 0.8 8

 Agriculture - livestock (manure) N2O-N 0.03

 Agriculture – soil management N2O-N 0.5

 Agriculture - field burning agricultural residues 0.001

 Fossil fuel combustion - transportation* 0.1

 Miscellaneous 0.1

NHx-N emissions2 3.1 31

 Agriculture: livestock NH3-N 1.6

 Agriculture: fertilizer NH3-N 0.9

 Agriculture: other NH3-N 0.1

 Fossil fuel combustion – transportation * 0.2

 Fossil fuel combustion - utility & industry * 0.03

 Other combustion 0.2

 Miscellaneous 0.1

 NOx-N emissions2 6.2 61

 Biogenic from soils 0.3

 Fossil fuel combustion – transportation * 3.5

 Fossil fuel combustion - utility & industry * 1.9

 Other combustion 0.4

 Miscellaneous 0.2

 Total Atmospheric inputs 10.0 100

Nr inputs to the Terrestrial environmental system

Atmospheric N depositionb 6.9 16

 Organic N3 2.1

 Inorganic NOy-N4 2.7

 Inorganic-NHx-N4 2.1

N fixation in cultivated croplands * 5 7.7 18

 Soybeans* 3.3

 Alfalfa* 2.1

 Other leguminous hay * 1.8

 Pasture* 0.5

 Dry beans, peas, lentils * 0.1

N fixation in non-cultivated vegetation * 6 6.4 15

N import in commodities *7 0.2 0.3

Synthetic N *8 15.1 35

 Fertilizer use on farms & non-farms 10.9

 Non-fertilizer uses 4.2

 Manure N production9 6.0 14

 Human waste N10 1.3 3

 Total Terrestrial inputs 43.5 100

Nr inputs to the Aquatic environmental system

 Surface water N flux11 4.8
13

Table 1 Notes

a. The Nr estimates in this table are shown with two 

significant digits or 0.1 million metric tons N per year 

(or Tg N/yr) to reflect their uncertainty; occasionally this 

report will show data to more significant digits, strictly 

for numerical accuracy. Because numbers are rounded, 

the sums of source category inputs do not always equal 

the total reported inputs. Obtaining quantitative estimates 

of each of the Nr terms and the associated uncertainties 

remains a major scientific challenge. 

b. Reducing the uncertainty in total deposition of 

atmospheric Nr to the surface of the 48 contiguous United 

States remains a scientific and policy priority. Based on 

observations and models, we estimate 5.9 (range 4 - 9) 

Tg N/yr total anthropogenic Nr deposition to the entire 

48 States. The EPA sponsored Community Multiscale Air 

Quality (CMAQ) Model run yielded a value of 4.8 Tg N/

yr. The value shown for the total (6.9 Tg N/yr) reflects the 

assumption that organo-nitrogen species should be added 

to the model estimate as 30% of the total. 

c. The synthetic N total includes 5.8 Tg N of fertilizer 

net imports to the United States (8.25 Tg N imported - 

2.41 Tg N exported) plus 9.4 Tg N of fertilizer produced 

in the United States in 2002.

Table 1 Data Sources (all data reflect N fluxes in the 

United States in 2002)

1 Emissions, N2O-N (U.S. EPA Inventory of U.S. 

Greenhouse Gas Emissions and Sinks, 1990-2006)

2 Emissions, NHx-N; Emissions, NOx-N (U.S. EPA 

National Emissions Inventory, release version October 

2007)

3 Atmospheric deposition, organic N (30% of total 

atmospheric N deposition, Neff et al. 2002)

4 Atmospheric deposition, inorganic NOy-N and 

NHx-N (U.S. EPA CMAQ model)

5 N2 fixation in cultivated croplands (USDA census of 

agriculture 2002, literature coefficients)

6 N2 fixation in non-cultivated vegetation (unpublished 

data estimate after Cleveland and Asner, 1990)

7 Net N imports in commodities and fertilizer trade 

(FAO - FAOSTAT)

8 Synthetic N fertilizer use (FAO - FAOSTAT and 

Association of American Plant Food Control Officials - 

AAPFCO)

9 Manure N production (USDA census of agriculture, 

literature coefficients)

10 Human waste N (U.S. Census Bureau population 

census, literature coefficients)

11 Surface water N flux (USGS SPARROW model, 

after Alexander et al., 2008)

2.2. Sources of New Nr to 

the Environment

Creation of “new” Nr in the environment refers to 

Nr that is either newly fixed within or transported into 

the United States. This “new” Nr highlights where Nr 

is introduced into ecosystems. New Nr arises from 

fossil fuel combustion, food production, and materials 

production (Table 1).

Fossil fuel combustion emits Nr (mostly NOx) to 

the atmosphere. Fossil fuel combustion introduces 3.5 

Tg N/yr and 1.9 Tg N/yr of NOx-N to the atmosphere 

from transportation, and utility/other industry sources, 

respectively (Table 1). Another 0.2 Tg N/yr of NH3-N and 

0.1 Tg N/yr of N2O-N is emitted from the same sources 

(Table 1). Thus the total amount of Nr created by fossil 

fuel combustion is 5.7 Tg N/yr, of which > 90% is in the 

form of NOx-N.

Synthetic Nr fertilizers are typically produced by the 

Haber-Bosch process and used primarily in agriculture 

to support food production. Production of fertilizers 

within the U.S. introduces Nr into U.S. terrestrial 

landscapes at the rate of 9.4 Tg N/yr, and net imports 

of fertilizer via world trade introduce 5.8 Tg N/yr. Of 

this total (15.2 Tg N/yr), 9.8 Tg N/yr is used as fertilizer 

on farms and 1.1 Tg N/yr is used on non-farms (i.e., 

residential and recreational turf-grass and gardens, and 

in explosives used by the mining industry), and 4.2 Tg 

N/yr is introduced for non-fertilizer uses, such as for 

production of plastics, fibers, resins, and for additives to 

animal feed (Table 1). 

Additional Nr is introduced into the U.S. from C-BNF

by agricultural legume crops such as soybean and alfalfa 

(7.7 Tg N/yr), and from imports of N contained in grain 

and meat (0.15 Tg N/yr) (Table 1). 

Thus in 2002, anthropogenic activities introduced a 

total of 29 Tg N into the U.S., mostly in support of food 

production, although turf production, industrial uses 

and fossil fuel combustion were also important sources. 

Natural sources of Nr in the U.S. are biological nitrogen 

fixation (BNF) in unmanaged landscapes, and lightning. 

The former contributes 6.4 Tg N/yr (Table 1) and the 

latter 0.1 Tg N/yr. Clearly, anthropogenic activities 

dominate the introduction of Nr into the U.S.

Losses of Nr to the environment in the U.S. occur 

during fossil fuel combustion and food production. The 

former occurs immediately, as Nr formation during 

combustion is inadvertent and the Nr, primarily as NOx, 

is emitted directly into the atmosphere. The latter occurs 

through all stages of food production and consumption. 

The following four Sections (2.2.1-2.2.4) of this report 

document the magnitude of the losses of Nr to the 

environment from the various components of both energy 

and food production.
14

2.2.1. Nr Formation and Losses to the 

Environment from Fossil Fuel Combustion

Fossil fuels such as coal, petroleum, and natural gas 

provide about 80% of all energy production in the U.S. 

(based on year 2000). When these fuels are burned at 

high temperatures, NOx is formed. The source of N is 

either the N contained in the fossil fuel or the N2 that 

makes up about 80% of atmosphere. Fuel-derived N is 

important in the case of burning coal (which contains 

N), while atmospheric-derived N2 is transformed to NOx

during higher temperature processes that occur when 

gasoline or diesel fuel is burned in motor vehicles (Table 

1). As Figure 3 indicates, in the U.S., highway motor 

vehicles account for the largest anthropogenic source of 

NOx (36%), followed by off-highway vehicles, electric 

utilities, and industrial processes. Emissions from aircraft 

make up only about 1% of the U.S. total for NOx, but a 

large proportion of this is released in free troposphere 

where lifetimes are long and adverse impacts wideranging. As such, continued reductions are encouraged 

(e.g., EPA Regulatory Announcement: New Emission 

Standards for New Commercial Aircraft Engines, www.

epa.gov/oms/regs/nonroad/aviation/420f05015.htm).

Figure 3 also illustrates that the amount of NOx (reported 

as metric tons of N) released from various fossil fuel sources 

has decreased dramatically from 1970. Total emissions were 

on the order of 7,400 metric tons in 1970 and decreased 

to 5,900 in 2002, with further decreases in 2006 to 5,030 

metric tons. Overall this represents a decrease of over 30%. 

The top sources (highway vehicles, off-highway vehicles, 

electric utilities, and other industrial and combustion 

systems) show decreases of 15-30% from 1990 to 2002 

(Figure 4). Reductions were the highest for “other” systems 

followed by electric utilities. These decreases are most likely 

the result of changes in regulations and control technologies 

for these stationary systems. More recent preliminary 

information provided by EPA indicates that electric power 

NOx emissions may have decreased 70% between 1990 and 

2009 and that the electric power sector’s NOx emissions 

now account for about 12% of anthropogenic NOx

emissions in the U.S. (U.S. EPA, 2010b). To a lesser extent, 

changes in highway vehicle regulations and the removal 

of older fleets from the road has resulted in a decrease of 

approximately 15%. This decrease however, is accompanied 

by an increase in miles traveled, which suggests that the 

actual decrease in a single vehicle is larger. Off-highway 

vehicles showed an increase in emissions, potentially due to 

better quantification of these sources. Such sources include 

locomotives and marine engines. EPA is in the process of 

implementing a number of regulations that will reduce NOx

emission from mobile sources (see Appendix F). Figure 

5 (provided by EPA) projects decreases in U.S. mobilesource NOx emissions. The implications of these recent 

regulations are not reflected in the quantitative analyses 

presented in this report. However, additional control of 

these and other sources could further decrease emissions. 

In fact, technological development in the locomotive 

industry shows that decreases of approximately 70% are 

possible. Further decreases would require more innovative, 

expensive methods such as selective catalytic reduction 

Data are reported as thousands of metric tons of N 

converted from NOx as NO2

Data source: www.epa.gov/ttn/chief/trends. More recent 

information provided by EPA indicates that electric power 

NOx emissions decreased 70% between 1990 and 2009 

(U.S. EPA, 2010b).

Figure 3: U.S. NOx emission trends, 1970-2006 Tg of Nitrogen (converted from short tons of Nox, as NO2)

Figure 4: Percent reductions in NOx emissions, 

1990-2002, from different sources 

(off-road, on-road, power generation, etc.)

Data source: www.epa.gov/ttn/chief/trends. More recent 

preliminary information (U.S. Environmental Protection 

Agency, 2010b) indicates that electric power NOx

emissions may have decreased 70% between 1990 and 

2009.
15

(SCR) with urea injection. Engine manufacturers are also 

investigating using SCR systems for diesels. However, it 

must be noted that these systems emit small amounts of 

NH3 and must be operated properly to avoid trading off 

NOx emissions for NH3 emissions. 

It should also be noted that it is difficult to control 

nitrogen emissions by regulating one source solely. As seen 

in Table 2 (data taken from 2001 for illustrative purposes 

have been reported with a higher degree of precision 

than is likely to be known), in Texas, fuel combustion 

sources are on the same order as emissions from highway 

vehicles. By comparison, in California, highway and 

off-highway vehicles are the dominant source of nitrogen 

emissions (over 75%). These results can be attributed to 

differences in the two states’ industries and power plants. 

In Texas,almost 40% of the power generation is from 

coal-fired plants. On the other hand, California imports 

most of its coal-fired power and generates its own power 

predominantly from other sources, such as natural gas 

(50%), hydro and nuclear (33%). Table 2 also shows 

emissions in Florida, Ohio, and Illinois. The emission of 

NOx from highway vehicles is likely related to population. 

For example, the estimated population of California in 

2006 was 36.4 million people versus Ohio and Illinois 

which are on the order of 11-12 million.

2.2.2. Nr Inputs and Losses to the Environment 

from Crop Agriculture

Agriculture uses more Nr and accounts for more 

Nr losses to the environment than any other economic 

sector. Synthetic fertilizers are the largest sources 

of Nr input to agricultural systems. The next largest 

source is N fixation in cultivated croplands (Table 1). 

The major pathways by which Nr is lost from these 

systems include NO3

-

 losses from leaching, runoff and 

erosion and gaseous emissions via volatilization of NH3

and NOx and nitrification/denitrification. Similar loss 

pathways occur for Nr that cycles through livestock 

systems, which also account for a large portion of Nr 

flux (predominantly as NH3) in animal agricultural 

systems (Aneja et al., 2006). Therefore, assessment of 

Nr impacts on the environment and development of 

strategies to minimize negative impact should be based 

on a thorough understanding and accurate accounting 

Figure 5: Mobile source NOx emission 

inventories

Source: Figure provided by Margaret Zawacki of the U.S. 

EPA Office of Transportation and Air Quality. Inventory 

data used to develop this figure are available in EPA’s final 

rule, Control of Emissions from New Marine CompressionIgnition Engines at or Above 30 Liters per Cylinder (www.

epa.gov/otaq/regs/nonroad/marine/ci/420r09019.pdf), 

except for onroad emissions, which EPA generated by 

running MOVES2010 (www.epa.gov/otaq/models/moves/

index.htm) at the national-month level.

Table 2: Examples of multiple sources from states with high NOx emissions based on 2001 data, 

and tons of NOx as NO2

TX CA FL OH IL

Fuel Combustion – Electric Util. 91,441 8,441 87,489 93,792 59,124

Fuel Combustion – Industrial 98,978 31,237 11,792 17,300 26,481

Fuel Combustion – Other 9,222 21,407 5,707 12,974 10,894

Industrial Processes 25,584 13,786 5,933 8,123 7,122

Highway Vehicles 164,937 182,471 116,889 83,593 78,278

Off-Highway Vehicles 106,162 85,064 38,475 46,239 52,797

Miscellaneous Sources 4,807 7,882 13,110 1,526 999

TOTAL, metric tons 501,151 350,301 279,778 263,561 235,817

Source: Data derived from the 2001 information obtained at: www.epa.gov/air/data/.

Mobile Source NOx Emission Inventories

0

2,000

4,000

6,000

8,000

10,000

12,000

14,000

16,000

18,000 Aircraft

Large commercial

marine

Locomotive

Nonroad diesel

Nonroad gasoline

On-road diesel

On-road gasoline

2002 2009 2014 2020 2030
16

of Nr fluxes in both crop and livestock systems, and 

the trends in management practices that have greatest 

influence on Nr loss to the environment from these 

systems (Aneja et al, 2008a,c).

As previously noted, in the past 60 years N fertilizers 

have had a beneficial effect on agriculture both nationally 

and globally by increasing crop yields. However, the high 

loading of Nr from agricultural nutrient sources has led to 

deleterious effects on the environment, such as decreased 

visibility from increased aerosol production and elevated 

N concentration in the atmosphere, ground, and surface 

waters (Galloway et al., 2003).

Nitrogen fertilizer use information

Obtaining accurate data on fertilizer use is a critical 

first step in understanding Nr cycles in agriculture. There 

are several sources of data reporting fertilizer usage 

but it is not clear whether data quality are sufficient 

for assessing environmental impact. Although the 

Uniform Fertilizer Tonnage Reporting System (UFTRS) 

was developed to collect fees to fund the consumer 

protection mission of state chemists and fertilizer 

regulatory control officials, it also provides data on 

fertilizer sales in many states, which in turn are used by 

many agencies and environmental scientists to estimate 

consumption and use of nitrogenous fertilizers in the 

U.S. The Association of American Plant Food Control 

Officials (AAPFCO ) tallies and publishes the statewide 

fertilizer sales data annually (Terry and Kirby, 2006). 

Information published by AAPFCO is a widely used 

source of fertilizer use data. It is typically assumed that 

fertilizers are used in the same region in which they 

were sold. The annual data published by AAPFCO , 

which are based on commercial fertilizer sold (and taxed 

in some but not all states), is the only state-level data 

source available. This state-level data source includes 

fertilizer sales for both agricultural and non-agricultural 

purposes. These state-level data must then be allocated 

to counties, regions, or watersheds in the states, and the 

algorithms used for this process are based on a number 

of assumptions that address dealer/farmer storage, 

inventories, and cross-state sales issues (personal 

communication, Stan Daberkow, USDA-ERS). 

The U.S. Department of Agriculture (USDA) 

National Agricultural Statistics Service (NASS) 

fertilizer usage data represents another source of 

information derived from farmer “agricultural chemical 

use” surveys that provide information in six categories: 

field crops, fruits and vegetables, nurseries/floriculture, 

livestock use, and post-harvest application. NASS 

periodically collects fertilizer, pesticide, and pest 

management data from a stratified random sample of 

farmers at the field level. The NASS report represents a 

useful data source but also would require extrapolation 

across reported crop acreage to represent a complete 

sample of application rates.

The UFTRS was not designed to track the source of 

inorganic nutrients applied to agricultural land on the 

geographic scale needed for watershed modeling. The 

system only tracks sales of synthetic fertilizers and not 

manure or biosolids applied to farmland. In addition, 

geographical data associated with each sale may or may 

not be near the actual point of application. However, 

given either regulatory or legislative changes (data 

reporting is mandated through each state’s fertilizer law), 

it could be possible to refine the current system used by 

each state’s department of agriculture to generate more 

precise data for improved modeling of watershed-scale 

nutrient mass balances. Those changes would help target 

interventions and extension programs to improve nutrient 

management and reduce nutrient losses. The lack of 

potential funding and the necessity to coordinate all the 

states involved limit the practicality of such an approach. 

State departments of agriculture have considered ways 

to improve the reporting system. Such improvements 

could include: 

1. An assessment to determine the needs for fertilizer 

usage data, the accuracy of the current data collection 

methods, and whether methods require revision to meet 

highest priority needs 

2. Improvements in the database format and web-based 

access

3. The identification of funding sources to support 

development of a more accurate, accessible, and 

comprehensive database system

4. Education and outreach to improve precision of 

reported fertilizer tonnage, including a clear distinction 

between nutrients used in crop, livestock, and nonagricultural operations 

In addition, the information could be refined to reflect 

site-specific data layers, although that would require 

development of a geospatial framework (and legal 

authority) to encourage reporting at the retail level where 

it is possible to collect geographic information.

The Chesapeake Bay watershed provides a good 

example of the fertilizer data dilemma. While the fertilizer 

tonnage that is currently being utilized to calibrate the 

Chesapeake Bay Program Watershed Model is relatively 

accurate, the county-specific tonnage may have an accuracy 

of only ±20 to 50% (Chesapeake Bay Scientific Technical 

Advisory Committee, 2007). For example, in a recent 

year, 17% of the reported tonnage was reported without an 

identified use, and there are indications some tonnage may 

have been reported more than once through the distribution 

chain (Chesapeake Bay Scientific Technical Advisory 

Committee, 2007). It is also possible that fertilizer reported 

for crop agriculture may actually have been used for lawn 

and turf, forestry, or other non-agriculture applications.

Nitrogen fertilizer application data collected on a 

specific crop-by-crop basis that can be associated with 
17

crop yields and location are essential for assessing both 

use patterns and efficiency. As previously mentioned, the 

USDA-NASS maintains a database on N fertilizer rates 

applied to the major crops (corn, wheat, cotton, soybeans, 

and occasionally other crops). The data are gathered from 

farmer surveys conducted every other year. These data 

represent a source of useful information (Protocols for 

Farming Reporting, Mark R. Miller, USDA, NASS). 

Data derived from NASS farmer surveys include six 

categories: field crops, fruits and vegetables, nurseries/

floriculture, livestock use, and post-harvest application. 

For each group, NASS collects fertilizer, pesticide, and 

pest management data on a stratified random sample of 

farmers at the field level. One field represents an entire 

farm for each sample in the field crops survey. Fruit and 

vegetable information are collected for the entire farm. If 

the field chosen for sampling has had manure applied in 

conjunction with inorganic fertilizer, only the inorganic 

portion will be reported because the survey does not ask 

about manure. 

Certain core crops are surveyed every other year on 

an even/odd basis and states are selected to cover at least 

80% of planted acres. There has been a recent five-year 

gap in the collection of NASS data for corn and other 

commodity crops. NASS was scheduled to resume 

coverage of corn and other commodity crops in 2010. 

This is a critical data gap and it is a problem given the 

large changes in corn price and production area during 

this period. Those data are needed to make progress in 

assessing fertilizer use and efficiency for major crops 

in the U.S. USDA NASS must resume their yearly data 

collection for commodity crops (Chemical Use Survey). 

Potential environmental impacts of increased N inputs 

associated with expanded corn acreage for biofuel 

production cannot be properly evaluated in the absence of 

such critical nutrient management data.

Based on the NASS survey data, the USDA Economic 

Research Service released a report on fertilizer use 

that provided data on fertilizer consumption and type 

of fertilizer used from 1960-2006 (Figure 6) and types 

of fertilizers used (Table 3) (USDA, 2008). The share 

of crop area receiving fertilizer and fertilizer use per 

receiving acre, by nutrient, are presented for the major 

producing states for corn, cotton, soybeans, and wheat. 

Additional data include fertilizer farm prices and indices 

of wholesale fertilizer price. 

Finding 1: Crop agriculture receives 60% of U.S. 

annual new Nr inputs from anthropogenic sources (9.8 

Tg from N fertilizer, 7.7 Tg from crop BNF versus 29 

Tg total) and accounts for 58% (7.6 Tg) of total U.S. 

Nr losses from terrestrial systems to air and aquatic 

ecosystems. However, current monitoring of fertilizer use 

statistics by federal agencies is inadequate to accurately 

track trends in quantities and fate of N applied to major 

crops and the geospatial pattern by major watersheds. 

Recommendation 1: The Committee recommends 

increasing the specificity and regularity of data 

acquisition for fertilizer application to major agricultural 

crops in terms of timing and at a sufficiently small 

application scale (and also for urban residential and 

recreational turf) by county (or watershed) to better 

inform decision-making about policies and mitigation 

options for reducing Nr load in these systems, and to 

facilitate monitoring and evaluation of impact from 

implemented policies and mitigation efforts.

Nitrogen fertilizer use efficiency

Nitrogen fertilizer use efficiency (NFUE) is critical 

because higher use efficiency leaves less N remaining 

to create potential environmental problems. Here and 

throughout this report we define NFUE as the grain 

yield per unit of applied N, which is the product of two 

parameters: (1) the proportion of applied N fertilizer that 

is taken up by the crop, or N fertilizer recovery efficiency 

[recovery efficiency (RE) in kg N uptake per kg N applied], 

and (2) the physiological efficiency with which the N

taken up by the crop is used to produce economic yield for 

crops such as grain or fruit [physiological efficiency (PE) 

in kg yield per kg N uptake] (Cassman et al., 2002)13. All 

else equal, when higher NFUE is achieved without yield 

reduction, the crop takes up more of the applied N and 

incorporates it into its biomass, which leaves less of the 

applied Nr at risk for losses via leaching, volatilization, or 

denitrification. Fixen (2005) reports that there is substantial 

opportunity for increasing NFUE through development 

Figure 6: Fertilizer consumption in the United 

States, 1960 to 2006 

Source: Slater et al., 2010. Reprinted with permission 

from the Association of American Plant Food Control 

Officials.

13 N fertilizer use efficiency (NFUE) is calculated as the ratio of grain yield to the quantity of applied N fertilizer (kg grain/kg applied N).
18

and adoption of more sophisticated nutrient management 

decision aids. 

In most cropping systems, RE is the most important 

determinant of NFUE. A recent review of RE for cereals 

based on field studies around the world, mostly conducted 

on “small-plot” experiments at research stations, reported 

mean single year RE values for maize, wheat and rice of 

65%, 57%, and 46%, respectively (Ladha et al., 2005). 

However, crop RE values based on actual measurements 

in production-scale fields are seldom greater than 50% 

and often less than 33%. For example, a review of RE in 

different cropping systems estimated average recoveries 

of 37% for maize in the north central U.S. (Cassman et 

al., 2002). It is also important to note that soil N provides 

the majority of the N taken up by most crops grown on 

soils with moderate to good soil fertility. For maize in the 

U.S. Corn Belt, for example, 45-77% of total N uptake 

was estimated to come from soil N reserves, based on 

experiments from research stations (Sawyer et al., 2006). 

Therefore, highest N efficiency and economic return on 

N inputs are achieved when the amount and timing of 

applied N is synchronized with the availability of soil 

N throughout the growing season to minimize both the 

quantity of N input required and the N losses from soil 

and applied N sources.

However, there are relatively few data that provide 

direct measurement of N fertilizer recoveries by major 

field crops under production-scale conditions. Reducing 

the uncertainty in estimates of N fertilizer RE is 

fundamental for prioritization of research and education 

investments, both in the public and private sectors. While 

management can substantially improve RE on average, 

in any given year weather will always be an uncontrolled 

factor that can significantly influence system efficiency. 

Weather can influence system efficiency through effects 

on crop growth vigor and ability to acquire applied 

nutrients and through losses of nutrients due to runoff, 

denitrification, and leaching that can occur in periods of 

excessive rainfall. 

Although total N fertilizer use in the U.S. has increased 

more slowly in the past two decades (Figure 6), yields 

of all major crops have continued to increase. Because 

crop yields are closely related to N uptake (Cassman et 

al., 2002), these trends imply a steady increase in NFUE

and reduced N losses to the environment because more 

of the applied N is held in crop biomass and harvested 

grain. Greater NFUE has resulted from two factors. 

The first factor is a steady improvement in the stress 

tolerance of corn hybrids (Duvick and Cassman, 1999) 

that increases crop growth rates and allows sowing at 

higher plant densities, which together accelerate the 

establishment of a vigorous root system to intercept and 

acquire available N in the soil profile. The second factor 

is the development and adoption of technologies that 

may improve the congruence between crop N demand 

and the N supply for indigenous soil resources and 

applied N. Examples of such technologies include soil 

testing for residual nitrate and adjusting N fertilizer rates 

accordingly, split N fertilizer applications, fertigation 

(the application of nutrients through irrigation systems), 

site-specific management, and new fertilizer formulations 

(e.g., controlled release, nitrification inhibitors). For 

maize, which receives the largest share of total N fertilizer 

in the U.S. (44% in 2005), NFUE decreased markedly 

in the 1960s because N fertilizer rates rose more quickly 

than maize yields. However, with recognition of negative 

impact from over-application of N and associated N

losses to the environment (especially with regard to water 

quality) investment in research and education to improve 

N fertilizer efficiency resulted in more than 50% increase 

in NFUE from 1974-76 to 2002-05 (Figure 7). Similar 

Table 3: Types and amount of nitrogen fertilizers used in the United States in 2002

Synthetic Nitrogen Fertilizers Tg N/year % of total

Other 0.21 2

Urea 2.21 20

N Solutions 2.55 23

Anhydrous NH3 2.88 26

Ammonium phosphates and N-P-K blends 2.28 21

Ammonium sulfate, aqua ammonia, ammonium 

nitrate, and other nitrate and ammonical N 

fertilizers

0.76 7

Total * 10.89 100

Data from Terry and Kirby, 2006

* Because of number rounding, the sum of individual percentages does not equal 100%.
19

improvements have been documented for rice production 

in Japan and for overall crop production in Canada.

Despite these steady improvements, current levels 

of N fertilizer uptake efficiency appear to be relatively 

low, although data from production-scale studies are 

few (Cassman et al., 2002). Most farmers do not use 

best management practices (BMPs) with regard to 

nitrogen fertilizer management. For example, a recent 

U.S. Department of Agriculture Economic Research 

Service (USDA-ERS) Agricultural Resources and 

Environmental Indicators (AREI) report indicates that a 

majority of farmers still apply N in the fall, which gives 

the lowest fertilizer uptake efficiency and highest Nr 

losses compared to application in spring or during the 

crop growth period (USDA, ERS, 2006). This situation 

suggests substantial potential for improvement in NFUE 

and an associated reduction in Nr losses from crop 

agriculture, especially for maize in the warmer portions 

of the Corn Belt and other southern and southeast areas 

where maize is grown. One potential development is 

the use of controlled release fertilizers that release N in 

congruence with crop demand during the growing season. 

Although such fertilizers are already in use on high value 

horticultural crops, they are currently too expensive for 

lower value commodity grains such as corn, rice, or 

wheat. Control of N release should result in higher NFUE 

where there is high risk for N losses in cereal systems 

that receive the total amount of applied N in one or two 

large doses. Production-scale field studies are needed 

to document the benefits of this and other innovative 

technologies to improve NFUE.

As producers have significantly increased yields 

in commodity crops over the past 25 years, it is 

questionable whether university recommendations for 

nutrient applications are still current. Many university 

recommendations are now 20 to 25 years old. As a 

corollary to this problem, numerous environmental 

models of nutrient pollution are still utilizing older yield 

estimates, which often underestimate crop nutrient uptake 

and overestimate nutrient losses (Burgholzer, 2007). 

A systematic effort needs to be made to update data on 

crop yields used to estimate nutrient losses. The concept 

of NFUE should be emphasized as a way to address the 

need to balance economic and environmental goals. In 

fact, the development and adoption of technologies that 

improve nitrogen fertilizer efficiency can contribute to 

more profitable cropping systems through a reduction 

in fertilizer costs. For example, average NFUE in the 

U.S. required 1.0 kg of applied N to produce 43 kg of 

grain yield in the 1974-1976 period, whereas that same 

amount of N produced 65 kg of grain in the 2003-2005 

period (data taken from Figure 7). This gain in efficiency 

means that it is possible to achieve the 2004 U.S. average 

corn yield of about 150 bushels per acre (9,444 kg/ha) 

with 144 lbs per acre (161 kg/ha) of applied N fertilizer 

(based on the most recent NFUE achieved by U.S. corn 

producers) versus about 200 lbs per acre (224 kg/ha) of 

N fertilizer at the 1980 efficiency level. 

Nitrogen costs have become extremely volatile, 

mirroring natural gas prices. In late 2008, N fertilizer 

prices were more than double the 2006-2007 N fertilizer 

prices. More recently, N fertilizer prices have fallen back 

to two thirds of the high following the decline of natural 

gas prices. If corn can be sold for $4.00 per bushel (25.5 

kg) and N costs $0.40 a pound (0.45 kg), this is a 10 

to 1 price ratio – the same as the $2.00 corn and $0.20 

nitrogen ratio that was typical from 2000 to 2005. There 

are also other critical factors in a farmer’s N application 

decisions, such as yield at the margin and weather. In the 

Corn Belt, one or two years in five may provide extremely 

Figure 7: Trends in corn grain produced per unit of applied fertilizer N (NFUE) in the United States

Source: Adapted from Fixen and West, 2002 (Figure 6). Adapted with permission; Copyright 2002, Springer 

Science+Business Media B.V. on behalf of the Royal Swedish Academy of Sciences.
20

favorable weather for corn production. A producer may 

view applying some extra N, hoping for good weather, 

as a reasonable economic gamble. If the yield response 

is more than half a bushel (12.7 kg) of corn per pound 

(0.45 kg) of N at the margin or if there is more than one 

extremely good year in five, the farmer benefits. 

Realistically, few farmers calculate their marginal 

returns from additional N in good years versus average, 

but the high corn-to-fertilizer price ratio encourages some 

farmers to plan for a good year and consider a larger N

application than might otherwise be appropriate for the N

utilization in the four years of lower yield. This presents a 

real dilemma if the policy goal is to reduce N transfers to 

the environment, especially in the four years of average 

or lower yields. Meeting this challenge will require 

approaches such as the development of real-time, inseason, decision-making tools that allow crop producers 

to use N fertilizer rates for average yields at planting and 

during early vegetative growth, and a final top-dressing 

as required to meet any additional N demand above this 

amount due to favorable climate and soil conditions 

that support higher than average yields (Cassman, 1999; 

Cassman et al., 2002). Robust crop simulation models 

using real-time climate data at a relatively localized 

geographic scale will be required to develop such tools.

Another option is to develop new, alternative crop 

production systems that require less N fertilizer. Such 

systems may employ legume cover crops, more diverse 

crop rotations, and tighter integration between crop and 

livestock production to achieve greater reliance on N

inputs from legume N fixation and recycling of N in 

manure and compost. At issue, however, is whether such 

systems actually reduce Nr losses to the environment 

because the same loss mechanisms and pathways operate 

on N from both commercial fertilizer and organic 

sources. Also at issue is the indirect land use change 

impact from widespread adoption of these more diverse 

cropping systems because they have reduced crop yields 

per unit land area compared to more simplified crop 

rotations such as corn-soybeans that receive N fertilizer. 

Lower yields would require more land in production to 

meet food demand. Therefore, a key issue is whether the 

tradeoff in reduced N fertilizer inputs to more diverse 

crop rotations with organic N inputs would actually 

result in less Nr losses to the environment compared to 

conventional cropping systems that require less land to 

produce the same amount of crop output. 

Another approach to reduce Nr losses from agriculture 

would be to shift and/or adjust cropping systems across 

the landscape. This would involve changes in land use as 

well as crops. This approach for parametric reductions 

in nitrogen was analyzed extensively in the Gulf of 

Mexico Hypoxia Assessment (Doering et al., 1999). As 

part of a modeling exercise to address opportunities and 

consequences for reducing Nr, crop rotations as well as 

tillage practices and fertilizer inputs were adjusted to 

meet successive constraints on excessive Nr while also 

maximizing consumer and producer welfare to the extent 

possible. The model favored those crops and cropping 

practices that had lower Nr leakage. Where the model 

could not find a crop production system at given locations 

that allowed positive net returns to the land, that land 

was taken out of production. At a 20 percent Nr reduction 

scenario, crop acreage was reduced by about 6%. This 

analysis was based on crop genetics, rotations, and tillage 

practices as of the 1990s. 

Unintended impacts of lower application rates of 

nitrogen for crop production

Crop production and environmental quality are 

potentially lost or gained at the expense of each other. 

Although leakage of N from crop production systems 

cannot be eliminated, N losses can be minimized 

substantially. One way to decrease leakage is to apply 

less N fertilizer to croplands. For example, Hu et al. 

(2007), using the SWAT model, predict that decreasing N

fertilizer application rates by 10 to 50% of those used in 

the 1990s in the upper Embrarras River watershed in east 

central Illinois would decrease NO3

-

 output to the river 

by 10 to 43%. This simple “solution” can cause problems 

for crop production as yields and crop quality (protein 

content) may decrease, causing economic loss to the 

farmer, decreased food quality for the consumer and, at a 

global scale, a reduction in food security. 

Cropping systems managed in a consistent manner 

over time reach a functional equilibrium between nitrogen 

inputs and outputs. Because crop yields are closely 

linked to the quantity of N accumulation in above ground 

biomass at maturity (Cassman et al., 2002), there would 

be a proportional decrease in crop yields in response 

to a decrease in the amount of N fertilizer application. 

The magnitude of this yield reduction would depend 

on the magnitude of decrease in the rate of applied N

and the efficiency of N uptake from the applied N, as 

well as interrelationships with the availability of other 

nutrients. Hence, yield reductions can be mitigated, or 

even eliminated, if methods and fertilizer formulations 

used in fertilizer-N application increase the efficiency of 

N uptake to offset the reduction in the amount of applied 

N. It is also important to note that reduced or insufficient 

N rates for crop production risk impairment of long-term 

soil productivity. Jaynes and Karlen (2005) reported that 

N rates below the agronomic and economic optimum 

could degrade the soil resource and decrease soil organic 

matter over time. Thus care must be exercised in any N

rate adjustments to protect soil productivity and to support 

soil resource sustainability.

Another concern associated with decreasing N

fertilizer input to cereal crop production is the effect on 

crop quality. In Denmark, for example, national efforts 

to decrease Nr losses to the environment resulted in a 

large reduction in use of N fertilizer in crop production 

(Figure 8). This situation occurred as part of Denmark’s 
21

response to the European Union Nitrate Directive. 

Synthetic fertilizer N use in Denmark decreased from 

approximately 0.4 Tg N in 1991 to 0.2 Tg N in 2002. 

Animal manure N application decreased from 0.25 Tg 

N to approximately 0.24 Tg N during this time period. 

Despite this large decrease in N fertilizer use, cereal 

crop yields remained relatively constant (data not 

shown). But while grain yields were maintained, there 

was a decrease in grain quality as determined by protein 

content (Figure 9). Grain protein content in wheat is 

critical for determining its quality for bread. In the 

U.S., a grain protein content of 12% is considered the 

threshold for good quality bread wheat, and N fertilizer 

application rate has a large influence on determining this 

trait (Cassman et al., 1992). As can be seen, grain protein 

content has declined from 12 to 10% in Denmark over 

the same period of lower fertilizer application rates.

Such trends raise several questions if declines 

continue or are found to be widespread. What is the cost 

to the farmer (considering human nutrition and end-use 

value costs)? Do these costs offset the environmental 

benefits created by decreasing N flows from crop 

production areas? What would be the regional and global 

impact if similar reductions in nitrogen fertilizer inputs 

to agriculture were put in place in developed countries 

that represent the largest source of grain exports to 

international markets?

As previously discussed, it is possible to increase 

crop yields without an increase in N fertilizer input if the 

methods used to apply N are modified to improve NFUE

as illustrated by U.S. maize yield trends and fertilizer 

use (Figure 7). However, at some point continued yield 

increases will require additional N input because of 

the tight relationship between yield and crop N uptake 

requirements. The goal of N fertilizer management 

should therefore be to achieve highest possible NFUE

while also sustaining crop yield increases adequate to 

meet food demand without need for a large expansion of 

crop production area. In fact, studies that evaluate NFUE

in production agriculture find a wide range in NFUE

among farmers. For farmers who achieve high NFUE, a 

reduction in N fertilizer rate would reduce yields. Hence, 

mandating an across-the-board reduction in N fertilizer 

use penalizes the good farmers who achieve high NFUE

as their yields would likely fall while it would have little 

effect on yields of farmers who currently over fertilize 

(Cassman et al., 2003; Dobermann and Cassman, 2004).

Finally, if yields were significantly reduced as a result 

of lower N fertilization rates, more land may need to 

be brought into production. Because nearly all prime 

agricultural land is already used for crop production, 

expansion of crop area will most likely occur on 

more marginal land, such as the land currently in the 

Conservation Reserve Program (CRP). Such conversion 

would have two negative results. First, additional N

losses from these acres would occur due to relatively 

low N fertilizer efficiency that typically occurs on 

marginal land that has multiple soil constraints to crop 

growth and yield. Second, conversion from CRP to crop 

production would result in loss of soil organic matter that 

accumulates under grassland, which in turn would lead to 

increased greenhouse gas emissions from agriculture. 

Finding 2: Nr inputs to crop systems are critical to 

sustain crop productivity and soil quality. Moreover, 

given limited land and water resources, global population 

growth, and rapid economic development in the world’s 

most populous countries, the challenge is to accelerate 

increases in crop yields on existing farm land while 

also achieving a substantial increase in N fertilizer 

uptake efficiency. This process is called “ecological 

intensification” because it recognizes the need to meet 

future food, feed, fiber, and energy demand of a growing 

human population while also protecting environmental 

Figure 8: Synthetic fertilizer and livestock manure 

N used as fertilizer in Denmark 1985-2003

Source: IFA, 2004. Reprinted with permission from the 

International Fertilizer Industry Association. Fertilizer N Used (Tg N)

Figure 9: Protein content of cereal grain 

in Denmark

Source: IFA, 2004. Reprinted with permission from the 

International Fertilizer Industry Association. 
22

quality and ecosystem services for future generations 

(Cassman, 1999). More diverse cropping systems with 

decreased Nr fertilizer input may also provide an option 

to increase NFUE on a large scale if the decrease in 

Nr losses per unit of crop production in these diverse 

systems can be achieved without a decrease in total food 

production, which would trigger indirect land use change 

to replace the lost production and negate the benefits. 

However, crop cultivars and agronomic practices are 

changing rapidly, which changes N requirements, but 

current efforts in research, extension, and conservation 

programs on N management within these rapidly 

evolving systems are not adequate to meet the challenge 

of providing better information to increase NFUE.

Recommendation 2: To obtain better information 

on Nr inputs and crop productivity, the Committee 

recommends that:

Recommendation 2a: Data on NFUE and N mass 

balance, based on direct measurements from productionscale fields, should be generated for the major crops 

to identify which cropping systems and regions are of 

greatest concern with regard to mitigation of Nr load and 

to better focus research investments, policy development, 

and prioritization of risk mitigation strategies.

Recommendation 2b: Efforts at USDA and 

universities should be promoted to: (i) investigate means 

to increase the rate of gain in crop yields on existing 

farm land while increasing N fertilizer uptake efficiency 

and (ii) explore the potential for more diverse cropping 

systems with lower N fertilizer input requirements so long 

as large-scale adoption of such systems would not cause 

indirect land use change.

Recommendation 2c: EPA should work closely with 

the USDA, Department of Energy, and the National 

Science Foundation, and universities to help identify 

research and education priorities to support more efficient 

use and better mitigation of Nr applied to agricultural 

systems. 

Biological fixation in cultivated croplands

Reactive nitrogen is also introduced to the landscape 

in significant quantities via BNF in cultivated crop 

lands. Management of biologically fixed N, insofar as 

it is possible, is proportionally as critical a task as the 

management of synthetic N because Nr from BNF is 

prone to the same loss pathways as Nr from commercial 

fertilizers. To quantify BNF due to human cultivation of 

crops, the Committee calculated the annual agricultural 

fixation for 2002 using crop areas and yields reported 

by the USDA Census of Agriculture (USDA, 2002). The 

Committee multiplied the area planted in leguminous 

crop species by the rate of N fixation specific to each 

crop type, assigning rates based on a literature review, 

as summarized in Table 4 below and shown relative 

to other inputs in Table 1. Annual nitrogen inputs to 

cropping system from BNF by legume crops was 7.7 

Tg N/yr in 2002, accounting for ~15% of the overall Nr 

inputs to the terrestrial landscape from all sources and 

20% of the agricultural sources (Table 1). Soybean and 

alfalfa contributions are the most important agricultural 

legumes in terms of nitrogen input and contribute 69% 

of total BNF inputs in U.S. agriculture. 

Emissions factors and losses to the environment 

from fertilizers and organic nitrogen sources 

Agriculture is a significant contributor of Nr inputs 

into the atmosphere. Nitrogen fertilizer losses vary greatly 

Table 4: Estimates of nitrogen input from biological nitrogen fixation from major legume crops, 

hay, and pasture

Nr fixation in cultivated croplands

production area, 

Mha rate, kg/ha/yr Tg N/yr % of total *

Soybeans 29.3 111 3.25 42

Alfalfa 9.16 224 2.05 27

Other leguminous hay 15.4 117 1.80 23

Western pasture 161 1 0.16 2

Eastern pasture 22.0 15 0.33 4

Dry beans, peas, lentils 0.88 90 0.08 1

Total 7.67 100

Source: Updated estimate for soybean based on a generalized relationship between soybean yield and the quantity of 

N fixation (Salvagiotti et al., 2008). Other values are from Boyer et al. (2002).

* Because of number rounding, the sum of individual percentages does not equal 100%.
23

due to differences in soil properties, climate, and the 

method, form, amount, timing and placement of applied N

(Cassman et al., 2002). In addition, any factor that affects 

crop growth vigor and root system function also affects 

the ability of the plant to recover applied N efficiently. 

For example, denitrification can range from 0% to 70% of 

applied N (Aulakh et al., 1992). This process is mediated 

by heterotrophic, facultative anaerobic soil bacteria that 

are most active under warm, wet soil conditions; they 

have low activity in dry sandy soils. 

Despite this variation, watershed, regional, and 

national assessments of carbon and N cycling often 

rely on average values for losses from each pathway. 

For example, the Intergovernmental Panel on Climate 

Change (IPCC ) assumes that 1% of applied N fertilizer 

(uncertainty range of 0.3-3.0%) is lost from direct 

emissions of N2O at the field level due to nitrification/

denitrification. This assumption is based on analysis 

of all appropriate scientific publications that report 

these losses for specific crops and cropping systems 

(IPCC , 2007a). The same 1% default emission factor 

for field-level N2O emission is applied to other N

inputs from crop residues, organic amendments such as 

manure, and from mineralization of native soil organic 

matter. Data from scores of field studies were used to 

obtain this average value. A number of recent studies 

confirm that N2O losses to the environment during the 

growing season at the field level can represent <1% 

of the applied nitrogen – even in intensive, high-yield 

cropping systems (Adviento-Borbe et al., 2006). Despite 

these average values, it is also clear that N2O losses 

can vary widely even within the same field and from 

year to year due to normal variation in climate and 

crop management (Parkin and Kaspar, 2006; Snyder 

et al., 2007). Moreover, the loss of nitrogen from 

agricultural watersheds is strongly dependent on climate 

change (e.g., rainfall changes). Predicted increases and 

decreases in rainfall will likely have a dramatic impact 

on nitrogen export from agricultural fields. For example, 

precipitation is predicted to increase in the upper 

Mississippi watershed and, other factors being equal, N

export should increase (e.g., Justic et al., 1995a,b).

Additional indirect N2O emissions result from 

denitrification of volatilized NH3 deposited elsewhere 

or from NO3

-

 lost to leaching and runoff as the Nr 

cascades through other ecosystems after leaving the 

field to which it was applied. Here the IPCC assessment 

protocol assumes that volatilization losses represent 10% 

of applied N, and that N2O emissions for these losses 

are 1% of this amount; leaching losses are assumed 

to be 30% of applied nitrogen and N2O emissions are 

0.75% of that amount (IPCC , 2007a). Therefore, the 

IPCC default value for total direct and indirect N2O

emissions represents about 1.4% of the applied N from 

fertilizer. By the same calculations, 1.4% of the N in 

applied organic matter, either as manure or compost or in 

recycled crop residues, is also assumed to be emitted as 

N2O. Recent work funded by EPA used the DAYCENT

model to estimate N2O emissions from cropping systems 

(Del Grosso et al., 2005). However, due to the cost of 

field validation of such models, there are relatively few 

validations across a representative range of cropping 

systems and environment. Therefore, it is not clear that 

use of such a complex model gives better estimates of 

N2O emissions than the more straightforward IPCC

assessment protocol.

Others have estimated higher average global N2O

losses of 3-5% of applied nitrogen fertilizer based on 

historical changes in atmospheric N2O content and 

changes in Nr production during the past 50 years 

(Crutzen et al., 2008), as opposed to the field-based 

estimates that form the basis of IPCC estimates. Because 

N2O is such a potent GHG, and given the more than 

two-fold difference in estimates of N2O losses, there is a 

critical need to improve understanding and prediction of 

N2O losses from agricultural systems. N2O emissions in 

the U.S. are estimated to be 0.78 Tg N/yr (Table 5) (U.S. 

EPA, 2005b).

Biogenic NOx emissions from croplands are on the 

order of 0.5% of fertilizer input – much more than this in 

sandy soils and less as clay content increases (Aneja et al. 

1996; Sullivan et al. 1996; Veldkamp and Keller, 1997; 

Civerolo and Dickerson, 1998). However, NOx emissions 

by agricultural burning are relatively unimportant. 

Ammonia volatilization of N from applied fertilizer 

can be the dominant pathway of N loss in rice soils and 

can account for 0% to more than 50% of the applied N

depending on water management, soil properties, and 

method of application (citations within Peoples et al., 

Table 5: N2O emissions in the United 

States, 2002

Tg N/yr % *

Agricultural Soil 

Management 0.54 69

Manure Management 0.03 4

Mobile Combustion 0.09 12

Stationary Combustion 0.03 4

Nitric & Adipic 

Acid Production 0.05 6

Wastewater Treatment 0.02 2

Other 0.02 2

Total 0.78 100

* Because of number rounding, the sum of individual 

percentages does not equal 100%.
24

1995, 2004). Ammonia volatilization can be of the same 

range in upland cropping systems, with largest losses 

occurring typically on alkaline soils (Peoples et al., 

1995, 2004). The IPCC (2007a) uses a value of 10% of 

synthetic fertilizer N application and 20% of manure N as 

estimates of average NH3 volatilization.

Taken together, N losses from all forms of direct 

gaseous emissions from crop production systems can 

represent a substantial portion of applied N fertilizer 

when soil conditions favor such emissions and there is a 

lack of synchrony between the amount of N applied and 

the immediate crop demand (Goulding, 2004). Therefore, 

achieving greater congruence between crop demand and 

the N supply from fertilizer is a key management tactic 

to reduce N losses from all sources. Success in reducing 

N losses and emissions from agriculture will depend on 

increased efforts in research and extension to close gaps 

in our understanding of N cycling and management in 

crop production, especially as systems further intensify 

to meet rapidly expanding demand for food, feed, fiber, 

and biofuel.

Finding 3: Nitrous oxide emissions from the Nr inputs 

to cropland from fertilizer, manure, and legume fixation 

represent a large proportion of agriculture’s contribution 

to GHG emissions, and the importance of this source of 

anthropogenic GHG will likely increase unless NFUE is 

markedly improved in crop production systems. Despite 

its importance, there is considerable uncertainty in the 

estimates of nitrous oxide emissions from fertilizer and 

research should focus on reducing this uncertainty. 

Recommendation 3: The Committee recommends 

that EPA ensure that the uncertainty in estimates of 

nitrous oxide emissions from crop agriculture be greatly 

reduced through the conduct of EPA research and through 

coordination of research efforts more generally with other 

agencies such as USDA, DOE, NSF and with research 

conducted at universities.

Impact of biofuel production capacity on Nr flux 

in agriculture

The enormous use of liquid fuels in the U.S., the 

rising demand for petroleum based liquid fuels from 

countries like China and India, and the decline in 

petroleum discovery all contributed to the recent record 

high petroleum prices. In addition, most of the world’s 

petroleum reserves are located in politically unstable 

areas. This has provided strong motivation for policies 

promoting investment in biofuels made from corn, 

oil crops, and ultimately from cellulosic materials. In 

the U.S., ethanol production capacity from corn more 

than doubled from 2006 to 2009 (to a capacity of over 

47 billion liters/year in January 2009). The renewable 

fuels standard in the 2007 Energy Independence and 

Security Act (EISA) will support another 9.5 billion 

liters/year of corn-based ethanol by 2015. An additional 

79.5 billion liters is to come from cellulosic ethanol 

by 2022. Production of biodiesel from vegetable oils 

also is encouraged in EISA, but expansion has been 

slowed by the high food value of such oils. Brazil 

is rapidly expanding its production of relatively low 

cost sugarcane ethanol and U.S. policies continue to 

be aimed at bringing about increased future biofuel 

production in the U.S.

In 2007 and 2008 petroleum prices pushed ethanol 

prices high enough to draw corn from food and feed 

uses into ethanol production and contribute to the 

increased price of corn. Because of the increase in 

petroleum/ethanol prices and the government subsidy 

for ethanol production, 30% of the corn crop was used 

to produce ethanol in 2008 (Abbott et. al., 2008). With 

the subsequent collapse in petroleum and ethanol prices, 

followed by corn prices, there has been unused capacity 

in the U.S. ethanol industry as the corn/ethanol price 

ratio made ethanol production uneconomic for some 

firms. However, EISA is likely to lead to the production 

of cellulosic materials and even some expanded corn 

production for biofuels once the U.S. gets beyond 

the current blending limit for ethanol (Doering and 

Tyner, 2009). The higher corn prices of 2007 and 2008 

resulted in more land being planted to corn and higher 

N fertilizer requirements. Corn area went from 78.41 

million acres (31.73 million hectares) in 2006-2007 to 

93.6 million acres (37.88 million hectares) in 2007-

2008. Reduction in soybean production accounted 

for 12.01 million acres (4.86 million hectares) of 

the corn expansion, and the remaining new acreage 

came primarily from reduced cotton acreage and from 

hayland and pasture. This strong response to high 

demand for biofuel feedstock has led to concern about 

increased pressure on the environment from biofuels. 

One important factor is the increased N necessary for 

growing corn and cellulosic materials (Robertson, 

et. al. 2008). Biofuels production and consumption 

will also result in NOx emissions to the atmosphere. 

Expansion of corn or cellulosic materials production 

into marginal lands can be even more problematic with 

respect to nutrient leaching and soil erosion. Changes 

in N fertilizer prices add uncertainty to the additional 

amounts of N that may ultimately be used in biofuel 

feedstock production. Production of large amounts of 

distillers grains co-product is also changing the way 

that livestock feed rations are formulated, which in turn 

could have an influence on the cycling of N in cattle 

manure (Klopfenstein et al., 2008).

In February 2010, EPA released its final rule for the 

expanded Renewable Fuels Standard (RFS2) regarding 

greenhouse gas emissions from various types of 

biofuels based on life-cycle analysis as required by the 

2007 EISA. Standards for corn ethanol and soybean 

biodiesel are based on studies that include data from 

large-scale production systems because these are the 

only biofuels currently produced on a large commercial 
25

scale. In contrast, life cycle analysis (LCA) standards 

for cellulosic and other advanced biofuels were based 

on data from “pilot- and bench-scale” studies, or in 

many cases on hypotheses and rough estimates. While 

these initial estimates meet requirements imposed 

by 2007 EISA, the science underpinning life cycle 

assessments of biofuel systems, including direct and 

indirect land use change, are relatively undeveloped 

and evolving rapidly.

Finding 4: Rapid expansion of biofuel production 

has the potential to increase N fertilizer use through 

expanding corn production and its associated N fertilizer 

inputs and extending cultivation for cellulosic materials 

that will also need N. Distillers grains are changing 

animal diets and affecting N recycling in livestock. 

Both have important consequences for effective future 

management of Nr, but current models and understanding 

of how expanded biofuel production will affect Nr inputs 

and outputs from agriculture are inadequate to guide 

policy.

Recommendation 4: EPA should work with USDA 

and universities to improve understanding and prediction 

of how expansion of biofuel production, as mandated by 

the 2007 EISA, will affect Nr inputs and outputs from 

agriculture and livestock systems. Rapid expansion 

of biofuel production has the potential to increase N 

fertilizer use through expansion of corn production area 

and associated N fertilizer inputs, and from extending 

cultivation of cellulosic materials that will also need 

N inputs. Current models and understanding are not 

adequate to guide policy on how to minimize impact of 

biofuel expansion on environmental concerns related 

to Nr.

2.2.3. Nr Inputs and Losses from 

Animal Agriculture

In the U.S., domestic animals produce 6.0 Tg N/yr in 

manure and are the largest source of atmospheric NH3-N

(1.6 Tg N/yr) (Table 1). Livestock also contribute to 

N2O-N emissions, though in much smaller proportions 

(~4% of total U.S. N2O-N emissions).

Trends in Animal Agriculture

While animal production has been increasing since 

World War II, this report emphasizes the period from 

1970 to 2006. The production of chicken broilers 

increased more than four fold from 1970 to 2006 

(Figure 10) and milk production increased by nearly 

60% in this time period (Figure 11). Turkey production 

doubled and pork production increased about 25%, 

while meat from cattle (beef and dairy) remained 

constant (Figure 10).

Another trend in animal production has been for fewer 

animals to produce more animal products. For example, 

the 60% increase in the amount of milk produced in 2006 

compared to 1970 required 25% fewer cows (Figures 11 

and 12). Animal inventories declined by 10% for beef 

brood cows from 36 million head in 1970 to 33 million 

head in 2006, and the inventory of breeder pigs and 

market hogs declined 8% from 673 million head to 625 

million head in the same period, even with similar or 

greater annual meat production. This trend resulted from 

Figure 10: Meat production from 1970 to 2006 

Source: USDA, NASS, 2007 – Census Reports. Data on 

cattle were not taken after 1999.

Figure 11: Milk production from 1970 to 2006

Source: USDA, NASS, 2007 – Census Reports
26

greater growth rates of animals producing more meat in a 

shorter amount of time. In 1970, broilers were slaughtered 

after 80 days on feed at 1.7 kg live weight, but by 2006 

the average weight was 2.5 kg after only 44 days on feed 

(USDA, NASS, 2007).

Another trend in animal agriculture has been the 

increased size and smaller number of animal operations, 

which results from the mechanization of agricultural 

practices and increased specialization. There were only 

7% as many swine operations and 11% as many dairy 

operations in 2006 as there were in 1970 (Figure 13). 

There were half as many beef operations in 2006 as in 

1970, but beef operations also expanded in size while 

smaller producers held jobs off the farm.

All of these trends show an increase in management 

and labor efficiency to produce a similar or greater 

amount of animal products. Also, because animal 

production is more concentrated on fewer farms with 

greater specialization, fewer crops are produced on those 

farms. As a result, it is increasingly common to have more 

manure nutrients produced on a livestock farm than can 

be used efficiently as fertilizer for crops on that farm. 

Therefore, unless the manure is applied over a larger 

crop area, the resulting over-application of manure on the 

livestock farm can reduce the subsequent efficiency of its 

utilization and result in greater nutrient losses.

Impact of livestock production trends on 

nitrogen use efficiency

The trends in livestock production have both positive 

and negative environmental impacts. One of the 

significant positive impacts is that with smaller animal 

inventories producing greater quantities of animal 

products, there is an improved efficiency of nitrogen 

utilization per product produced. This effect is partly the 

result of effectively reducing maintenance requirements 

during production. The requirements for feeding animals 

can be divided into two components: maintenance and 

production. The maintenance component is the feed 

that is used to keep the animal alive and healthy so 

that production is possible. The production component 

includes the feed that is converted to animal protein and 

waste due to the inefficiencies of these conversions. The 

maintenance component depends upon the number of 

animals, each animal’s mass, and the time the animal is 

on feed. Thus, the maintenance requirement is diluted by 

faster growth rates and greater body weight at slaughter. 

The increases in production rates over time have led 

to greater efficiencies in N and P utilization for animal 

production and lower amounts of nutrients excreted per 

unit of animal protein produced.

Public concerns about the potential environmental and 

health effect of air emissions from CAFOs expand the 

impacts of food production beyond those associated with 

traditional agricultural practices (NRC, 2001; Aneja et al., 

2009). Increased emissions of N compounds from animal 

agriculture into the atmosphere may lead to increased 

odor and interactions in atmospheric reactions (e.g., 

gas-to-particle conversion) (Baek et al., 2004a; Baek and 

Aneja, 2004b). These compounds are then transported 

by wind and returned to the surface by wet and dry 

deposition processes, which may have adverse effects on 

human health and the environment (McMurry et. al, 2004; 

Aneja et. al, 2006b, 2008a,b,c; Galloway et. al, 2008).

Figure 12: U.S. inventory of mature dairy cows 

and milk production per cow from 1970 to 2006 

Source: USDA, NASS, 2007 – Census Reports

Figure 13: Number of animal operations in the 

United States from 1970 to 2006 

Source: USDA, NASS, 2007 – Census Reports
27

Adverse effects, further discussed in Chapter 3 of 

this report, include eutrophication, soil acidification, 

loss of biodiversity, and reactions that increase the 

mass concentration of atmospheric aerosols (PM2.5). 

Aerosol formation occurs when HNO3 reacts with basic 

compounds, and NH3 reacts with compounds. Ecosystem 

acidification can occur when HNO3 is deposited from the 

atmosphere. In addition, acidification can also occur when 

NHx is deposited due to the production of HNO3 from 

nitrification via soil microbes. Soil acidification occurs 

when HNO3 or NH4

+ deposits on soils with low buffering 

capacity and this can cause growth limitations to sensitive 

plant species. Deposition of NO3

-

 or NH4

+ also causes 

eutrophication (i.e., an over-abundance of nutrients), 

which can promote harmful algal growth leading to the 

decline of aquatic species. In fact, volatized NO3

-

 can 

travel hundreds of miles from its source, affecting local 

and regional biodiversity far from its origin (Aneja et al. 

2008b; James, 2008).

The potential for reduced environmental impact from 

Nr in livestock systems depends on the proportion of 

the total intake attributable to maintenance costs. The 

commonly used tables for diet formulation published 

periodically by the NRC for various animal commodities 

can be used to track diet formulation practices and 

assumptions regarding maintenance and production 

requirements. About one third of the energy intake 

recommended for growing broilers was assumed to 

be needed for maintenance (NRC, 1994) but protein 

requirements were not divided between maintenance and 

production. For example, a dairy cow producing 40 kg 

milk per year would divert about 25% of its energy and 

12% of its protein to maintenance (NRC, 2001).

In terms of nutritional efficiency of a herd or flock, 

maintenance of a productive phase (e.g., growth, 

lactation) also requires maintenance of a reproductive 

phase of the animal’s life cycle. In other words, the actual 

nutritional maintenance cost of a herd or flock is greater 

than it is for productive individuals only. For example, 

milk production requires non-lactating cows and heifers 

in the herd which do not produce milk but which consume 

nutrients. These additional maintenance costs are lower 

for broiler flocks than for cattle.

Finding 5: There are no nationwide monitoring 

networks in the United States to quantify agricultural 

emissions of greenhouse gases, NO, N2O, reduced sulfur 

compounds, volatile organic compounds (VOCs), and 

NH3. Satellite observations of agricultural emissions 

hold promise for providing a high degree of spatial 

coverage and may complement surface observations 

in this network, although robust methods are yet to be 

developed. In contrast there is a large network in place to 

assess the changes in the chemical climate of the United 

States associated with fossil fuel energy production, 

i.e., the National Atmospheric Deposition Program/

National Trends Network (NADP/NTN), which has been 

monitoring the wet deposition of sulfate (SO4

2-), NO3-

, 

and NH4

+ since 1978.

Recommendation 5: The status and trends of 

gases and particulate matter precursors emitted from 

agricultural emissions, e.g., NO3

-

 and NH4

+ should be 

monitored and assessed utilizing a nationwide network 

of monitoring stations. EPA should coordinate and 

inform its regulatory monitoring and management of 

reactive nitrogen with the multiple efforts of all agencies 

including those of the U.S. Department of Agriculture and 

NSF supported efforts such as the National Ecological 

Observatory Network (NEON) and the Long Term 

Ecological Research Network (LTER).

Changes in feeding practices

From 1970 to 2006, several feeding practices were 

changed for diets fed to livestock (NRC, 1977, 1988, 

1996, 2001). In 1989 and 1996, the NRC introduced the 

idea of dividing the form of protein fed to ruminants into 

that which is degraded by rumen microorganisms and that 

which passes through the rumen to be digested directly 

in the stomach and small intestine. Feeding ruminants 

with attention to rumen degraded and rumen undegraded 

protein decreases the amount of protein fed by 10 to 15% 

for a given protein requirement. For poultry and swine, 

manufactured amino acids were added to diets, decreasing 

the need for protein by 30%. Today, two amino acids 

(lysine and methionine), coated in a way to prevent 

degradation in the rumen, are sometimes added to dairy 

cattle diets thereby decreasing protein intake by another 

15% (NRC, 2001). Phytase added to swine and poultry 

diets in the past decade has decreased phosphorus feeding 

by 20 to 50% with some of the decrease attributed simply 

to better understanding phosphorus requirements.

It is difficult to estimate the combined effects 

of these changes in feeding practices, but we have 

calculated changes in manure N by using USDA NASS 

data and assuming improvements in both production 

rates and ration formulation (Table 6). In the case of 

beef cattle diet formulation, the changes in feeding 

practices were determined by comparing the NRC 1976 

recommendations with the NRC 1996 recommendations. 

Surprisingly, NRC 1996 recommended greater total crude 

protein compared to NRC 1976 despite formulating for 

rumen degraded and un-degraded protein and considering 

amino acid content. Therefore, improved diet formulation 

did not decrease N intake for beef in this time range but 

the effect of reduced maintenance did improve efficiency 

of N utilization.

Reduced nitrogen excretion from 

increased efficiency

Nitrogen excretion as fraction of animal production 

decreased from 1970 to 2006 (Table 6). However, in cases 

where the total amount of animal production in the U.S. 

increased substantially (e.g., broilers), total N excretion 

increased. The decrease in N excretion per unit of animal 
28

productivity was estimated by calculating the effects of 

changes in feeding practices and reduction of maintenance 

as described previously. The data in Table 6 indicate that 

there has been an increase in N utilization efficiency for 

livestock products.

For broilers, data are available to more accurately 

estimate the effect of changes in feeding and genetics 

on N excretion over time. However, these data do not 

represent the time period of interest in this report. 

Havenstein et al. (1994) compared a 1957 strain of 

broiler fed a 1957 diet to a 1991 strain fed a 1991 diet. 

Based on the reported N intake and production data, 

there was a 51% reduction in N excreted between these 

diets (Kohn, 2004).

Similarly, Kohn (2004) compared N excreted by U.S. 

dairy cows in 1944 and 2001. In 1944, the historically 

largest herd of dairy cattle in the U.S. (25 million cows) 

produced an average of 7 kg milk per cow per day (USDA

NASS, 2007). In 2001, nine million cows produced an 

average of 27 kg milk per cow per day. Assuming the 

cows in 1944 and 2001 were fed according to popular 

feeding recommendations of the time, the N intakes were 

360 and 490 g/day per cow, and N excretion rates (N

intake minus N in milk) were 326 and 364 g/day per cow. 

Multiplying by the number of cows in the U.S., shows 

that total milk production increased 40% from 52 billion 

kg to 73 billion kg, while N excretion decreased 60% 

from 3.0 Tg N to 1.2 Tg N, respectively.

Table 7 provides information on manure production 

from animal husbandry in the U.S. For Table 7, manure 

N was calculated for all U.S. animal agriculture using 

data on animal production from the 2002 Census of 

Agriculture (USDA, 2002). For data on livestock 

production (cattle, calves, poultry, hogs, and pigs), 

manure was calculated by the methods of Moffit and 

Lander (1999), following the exact methods they had used 

to compute manure from the 1997 Census of Agriculture, 

but using the updated information from the 2002 Census 

of Agriculture. For data on production of manure from 

other animals (horses, goats, and sheep), the table uses 

coefficients for manure excretion as a function of average 

animal weights and animal inventory, taken from Battye 

et al. (1994).

Volatilization of animal waste

Ammonia volatilization is highly variable and is 

influenced by the amount of total ammonical nitrogen 

(TAN), temperature, wind speed, pH, chemical and 

microbiological activities, diffusive and convective 

transport in the manure, and gas phase resistance in the 

boundary layer above the source (Arogo et al., 2006). 

EPA estimates annual manure N excreted in 

livestock production in the U.S. for the Inventory of 

U.S. Greenhouse Gas Emissions and Sinks (U.S. EPA, 

2007e). For the year 2002, these estimates indicate that 

a total of 6.8 Tg of N was excreted in livestock manure. 

Only a fraction of this N (~1.24 Tg) was recovered and 

applied directly as a nutrient source for crop production. 

Approximately 1.8 Tg N was transferred from the 

manure management systems, most likely by ammonia 

volatilization. Other loss vectors include leaching and 

runoff during treatment, and storage and transport before 

soil application. The remainder of the N was deposited 

in pastures and rangeland or in paddocks. This N is also 

Table 6: Livestock N excretion per kg production (g/kg) and per total United States (Tg/yr)

Commodity* 1970 2006

g/kg product Total United 

States g/kg product Total United States

Milk 17 0.89 11 0.92

Pork, live weight 57 0.56 42 0.54

Broilers, live weight 56 0.26 46 1.00

Beef, live weight 123 1.2 110 1.3

*Does not include manure produced for reproduction of stock (e.g., growing dairy heifers, breeder pigs).

Table 7: Manure production from animal 

husbandry in the continental United States, 

Tg N per year 2002.

Tg N/yr %

Cattle & Calves 4.35 72

Poultry 0.94 16

Hogs & Pigs 0.53 9

Horses, Goats & Sheep 0.19 3

Continental United States* 6.02 100

* Because of number rounding, the sum of manure 

production by category does not equal the reported total 

for the continental U.S.
29

susceptible to movement into the atmosphere and aquatic 

systems or incorporation into soil organic matter. By a 

combination of BMPs and engineered solutions it may be 

possible to reduce the emissions and discharge of odors, 

pathogens, and nitrogen compounds from agricultural 

operations (Aneja et al., 2008b,d).

Total manure production reported in Table 7 in the 

contiguous U.S. was estimated using USDA’s method 

and yields an estimate of 6.0 Tg N/yr; while EPA’s 

GHG inventory method in Table 8 yields a total for 

the U.S. of 6.8 Tg N/yr in 2002. The “greenhouse gas” 

method suggests 13% higher manure N production. This 

difference highlights uncertainty in the calculations. 

Although the values in Table 8 include Alaska and Hawaii, 

the values in Table 7 do not; however, given the relatively 

small amount of livestock production in those states, that 

does not contribute substantially to the difference.

Finding 6: Farm-level improvements in manure 

management can substantially reduce Nr load and 

transfer. While the NPDE S permitting process for 

CAFOs does include nutrient and manure management 

plans to limit the transport of land applied nutrients 

off farms, substantial off-farm transport of Nr still 

occurs, especially via atmospheric transport. There are 

currently very few incentives or regulations to decrease 

these transfers and loads despite the existence of 

management options to mitigate. 

Recommendation 6: A policy, regulatory, and 

incentive framework is needed and should be 

developed to improve manure management to reduce 

Nr load and ammonia transfer, taking into account 

phosphorus load issues.

2.2.4. Nr Inputs to Residential and Recreational 

Turf Systems

Turf grasses cover 12.6-16.2 million ha across the 

continental U.S. (Milesi et al., 2005). The area under 

turf grass is roughly the size of the New England states 

and occupies an area up to three times larger than that of 

irrigated corn (The Lawn Institute, 2007). The majority 

of this turf area (approximately 75%) is in residential 

lawns. About 80% of all U.S. households have private 

lawns (Templeton et al. 1998) that average 0.08 ha in size 

(Vinlove and Torla, 1995). Another approximately15 % 

of total turf grass area is in low-maintenance parks and 

approximately 10% is in athletic fields and golf courses, 

which often receive higher levels of N application due to 

hard use conditions. 

Supplemental N fertilization is often necessary to 

maintain healthy and aesthetically pleasing turf color, 

high shoot density, and the ability to resist and recover 

from stress and damage. Nitrogen also may be derived 

from atmospheric deposition or recycled decomposition 

of soil and grass clipping organic matter. Whether these 

inputs are sufficient to maintain lawns of adequate quality 

depends on many factors including age of the turf, uses, 

and expectations or goals of the homeowner or field 

manager. Also, turf grasses are used to stabilize soil, 

often with an erosion-prevention matrix such as organic 

mats or with hydroseeding. Depending on circumstances, 

these turf uses may be temporary until natural vegetation 

succeeds the turf, or may be low-maintenance turfs that 

are seldom fertilized, such as highway medians and 

shoulders, grassy swales and buffers.

Turf grass is maintained under a variety of conditions. 

Approximately 50% of all turf grass is not fertilized; 

the remainder is fertilized at varied intensities (A.M. 

Petrovic, personal communication, June 5, 2007). The 

Committee has arrayed the different turf managements 

into the following three groups according to the estimated 

amount of N-fertilizer applied annually (Table 9): 

residential lawns maintained by homeowners (0.73 

kg/100 m2), residential lawns cared for by professional 

lawn care companies (2.92 [range, 1.95-7.3] kg/100 m2), 

and athletic fields and golf courses (3.89 [range, 2.64-

6.64] kg/100 m2). The estimate of total N-fertilizer used 

Table 8: Fate of livestock manure nitrogen (Tg N)

Activity 1990 1992 1994 1996 1998 2000 2002 2004

Managed manure N applied to 

major crops 1.1 1.2 1. 2 1.2 1.2 1.3 1.2 1.3 

Manure N transferred from 

management systems 1.5 1.6 1.6 1.6 1.7 1.7 1.8 1.7 

Pasture, range, & paddock 

manure N 3.9 4.0 4.1 4.2 3.9 3.8 3.8 3.7

Total * 6.6 6.7 6.9 7.0 6.9 6.8 6.8 6.7

Source: U.S. EPA, 2007e

* Because of number rounding, the sums of manure production do not always equal the reported totals.
30

on turf grass in the U.S. is 1.1 Tg N/year, or 9% of the 

total average annual N-fertilizer used between 1999 and 

2005. Depending on land use patterns, certain areas of the 

country, particularly coastal areas where residential and 

urban properties prevail, turf fertilizer can be an important 

or even dominant source of nitrogen to surface waters.

Turf fertilizer N is susceptible to losses to the 

atmosphere and surface and ground water when it is not 

properly managed. Research on lawns has shown that 

leaching of NO3

-

 can range between 0 and 50% of N

applied (Petrovic, 1990). Nitrogen leaching losses can 

be greatly decreased by irrigating lightly and frequently, 

using multiple and light applications of fertilizers, 

fertilizing at the appropriate times (especially not too 

late in the growing season), and using soil tests to ensure 

proper balance of non-N soil condition and pH. In a soil 

column experiment with turf coverage, the percentage 

of N leached (as percentage of N applied) varied from 

8% to 14% using light irrigation and from 2% to 37% 

with heavy irrigation. Applying fertilizer in appropriate 

amounts, avoiding periods when grass is dormant, and 

not fertilizing too soon before irrigation or large rainfall 

events can all help ensure leaching and runoff will be 

minimal without affecting turfgrass color and growth 

(Mangiafico and Guillard, 2006). 

Nitrogen runoff losses are poorly quantified but a range 

similar to leaching is probable (A.M. Petrovic, personal 

communication). The chemical form of fertilizer N does 

not impact leaching/runoff unless the fertilizer is applied in 

late autumn (Petrovic, 2004a), although use of slow release 

or organic fertilizers can help reduce runoff and leaching. 

Shuman (2002) notes that runoff can be limited by applying 

minimum amounts of irrigation following fertilizer 

application and avoiding application before intense rain or 

when soil is wet. Losses of Nr to the atmosphere can be 

significant when urea is applied. Measured denitrification 

losses are usually small, but depend upon timing of N

application relative to soil water status, irrigation, and 

temperature. Typically 25% of N applied is not accounted 

for in runoff, leaching, and uptake/removal, or soil 

sequestration (A.M. Petrovic, personal communication). 

This suggests that volatilization and denitrification are 

important loss vectors. Nitrogen volatilization (Beard and 

Kenna, 2008) rates range from 0.9% under light irrigation 

to 2.3% under heavy irrigation.

While under-fertilization can lead to reduced grass 

stand and weed encroachment which results in more 

leaching and runoff N losses than from well managed 

lawns (Petrovic and Larsson-Kovach, 1996; Petrovic, 

2004b), Guillard (2008) recommends not fertilizing 

lawns of acceptable appearance. Further, prudent 

fertilization practices may include using one-third to 

one-half (or less) of the recommended application 

rate (i.e., application rates below 0.5 kg/100m2) and 

monitoring response (Guillard, 2008). Less or no 

fertilizer may produce acceptable lawns, especially once 

the lawn has matured, provided clippings are returned 

and mowing length is left high. 

As noted above, according to Petrovic (personal 

communication) half the lawns in the U.S. may not 

receive any fertilizer. Those lawns are presumably 

satisfactory to their owners. Further N reductions can 

be made if white clover is incorporated into turf, and 

grasses such as fescues (which require little or no N

supplements once mature) are selected for amenable 

parts of the country. These practices can potentially 

reduce N fertilization (and subsequent leaching risk) on 

turf by one third or more, saving 0.4 or more Tg N/year. 

When properly managed, turf grass provides a variety 

of services that include decreasing runoff, sequestering 

carbon dioxide, and providing a comfortable environment 

in which to live (Beard and Green, 1994). 

Finding 7: Synthetic N fertilizer application to urban 

gardens and lawns amounts to approximately 9% of the 

total annual synthetic N fertilizer used in the United 

States. Even though this N represents a substantial portion 

of total N fertilizer use, the efficiency with which it is 

used receives relatively little attention.

Recommendation 7a: To ensure that urban 

fertilizer is used as efficiently as possible, the Committee 

Table 9: Estimate of fertilizer N used on turf grass in the United States in the year 2000, 

based on a total area of 12.6 million ha

Type of Turf Fertilized Area (Million ha) N rate (kg/ha/yr)* Total N Used (Tg/yr) 

Nominal Fertilization 4.7 73 0.35

Professional Lawn Care 0.93 296 (195-488) 0.27

High Maintenance Areas 

(golf/sports) 1.26 390 0.49

Total 6.89 – 1.11

Source: Data derived from Milesi; et al., 2005 and A.M. Petronic, Personal Communication 2007

*A conversion factor of 1000 m2/ha was used with application rates (kg N/100 m2) of: 0.73 for nominal fertilization, 2.92 

for professional lawn care, and 3.89 for high maintenance areas kg N/100 m2
31

recommends that EPA work with other agencies such as 

USDA as well as state and local extension organizations 

to coordinate research to ensure that fertilization 

recommendations are accurate and promote awareness of 

the issue. 

Recommendation 7b: Through outreach and 

education, supported by research, improved turf 

management practices should be promoted, including 

improved fertilizer application and formulation 

technologies and maintenance techniques that minimize 

supplemental Nr needs and losses, use of alternative turf 

varieties that require less fertilization, alternative ground 

covers in place of turf, and use of naturalistic landscaping 

that focuses on native species. 

2.3. Nr Transfer and Transformations in 

and Between Environmental Systems 

This section discusses the transfers and flows of Nr 

within and between environmental systems (ES), which 

include atmospheric, terrestrial, and aquatic environments. 

The first Section (2.3.1) contains information on Nr 

deposition from the atmosphere to terrestrial and aquatic 

systems, presents estimates of input and recycling of 

Nr within terrestrial systems, and discusses movement 

of Nr from the terrestrial to the aquatic system. The 

second Section (2.3.2) presents an estimate of storage of 

Nr within the terrestrial system. Areas of uncertainty in 

Nr transfer and transformation are discussed in Section 

2.3.3. Section 2.3.3 also contains an example analysis of 

Nr input and fate in 16 watersheds in the northeast U.S. 

No comprehensive national data are available to assess 

the transfer and transformations of Nr in and between 

the atmosphere, terrestrial systems and aquatic systems. 

The example analysis in Section 2.3.3 shows how an 

evaluation of inputs and fate of Nr could be conducted for 

a large watershed.

2.3.1. Input and Transfers of Nr in the 

United States 

Nitrogen deposition from the atmosphere to the 

earth’s surface

Atmospheric input contributes substantially to the Nr 

content of terrestrial and aquatic ecosystems for the U.S., 

but the magnitude and mechanisms of Nr deposition to the 

earth’s surface remain major unanswered environmental 

questions. Along the eastern U.S. coast and eastern 

Gulf of Mexico, atmospheric deposition of N currently 

accounts for 10% to over 40% of new N loading to 

estuaries (Paerl et al., 2002). Other watershed contribution 

estimates range widely throughout the U.S., depending 

on size of the watershed, the size of the estuary, and the 

magnitude of contributing sources of atmospheric N

enrichment. Valigura et al. (2001) identified a median 

atmospheric N contribution of about 15% for 42 

watersheds located throughout the U.S., although the 

maximum estimate was 60%. 

NOx, NH3 and their reaction products not deposited 

onto continents are generally lofted into the free 

troposphere where they can have a wide range 

of influence and, in the case of NOx, because of 

nonlinearities in the photochemistry, generate substantial 

amounts of tropospheric ozone (U.S. EPA, 2006a). 

Total N deposition involves both gases and particles, 

and both dry and wet (in precipitation) processes. Rates 

of deposition for a given species (in units of mass of N

per unit area per unit time) can be measured directly, 

inferred from mass balance of the atmospheric budget, 

or modeled numerically, but substantial uncertainties 

remain with each of these techniques when applied 

to deposition of any Nr species. A portion of the Nr 

deposited to the earth’s surface is re-emitted as NH3, 

NO, or N2O (Galbally and Roy, 1978; Kim et al., 1994; 

Civerolo and Dickerson, 1998; IPCC, 2007a; Crutzen et 

al., 2008). Although naturally-produced Nr is involved, 

anthropogenic Nr dominates over most of the U.S. In 

Appendix A we provide a review of the state of the 

science concerning the total annual Nr deposition and 

trends in that deposition to the contiguous 48 states.

Deposition involves both oxidized and reduced N

species. Of the oxidized forms of atmospheric N, all the 

members of the NOy family (NO, NO2, NO3, N2O5, 

HONO, HNO3, NO3

-

, PAN and other organo-nitrates, 

RONO2) can be transferred from the troposphere to the 

surface, and some (e.g., NO) undergo bidirectional flux. 

Volatile amines are also detected as NOy compounds 

(Kashihira et al., 1982; Wyers et al., 1993). Although 

a potent GHG, N2O is only emitted, not deposited 

and therefore has not been considered in the material 

presented here and in Appendix A. Of the reduced forms 

of atmospheric nitrogen, NH3 and NH4

+ play a major 

role. There is also evidence of deposition of organic N

such as amino acids and isoprene nitrates, and recent 

observations suggest that these can account for ~10% 

(possibly as much as 30%) of the U.S. NOx budget, 

especially in summer (Keene et al., 2002; Horowitz et al., 

2007; Duce et al., 2008; Sommariva et al., 2008). While 

this is a worthy research topic, measurements are still 

limited and deposition of organic N compounds has not 

been reviewed in this report. The wide array of relevant 

atmospheric compounds makes direct measurement and 

accurate load quantification challenging. Appendix A 

provides a discussion of wet and dry deposition of Nr in 

the U.S. and the relationship between emissions of Nr and 

observed deposition.

Consideration of NOy as a supplement for 

the current NO2 National Ambient Air Quality 

Standard

The six principal (or criteria) pollutants for which 

EPA has established National Ambient Air Quality 

Standards (NAAQS) include “oxides of nitrogen” (the 

sum of NO and NO2 ) or NOx. The specific chemical 

compound nitrogen dioxide (NO2) has been selected as 
32

the indicator for compliance with the NAAQS for NOx. 

The levels of primary and secondary standards for NO2

are identical at 0.053 ppm (approximately 100 μg/m3) in 

annual arithmetic average, calculated from the one-hour 

NO2 concentrations. In forming an integrated policy 

for protecting the environment from adverse effects of 

reactive nitrogen, it is appropriate to consider whether 

the existing criteria pollutants are sufficiently inclusive of 

Nr species. EPA’s recent Integrated Science Assessment 

for Oxides of Nitrogen and Sulfur – Ecological Criteria 

(U.S. EPA, 2008d) evaluated the scientific foundation for 

the review of the secondary (welfare-based) NAAQS for 

oxides of nitrogen and concluded that: 

The instrumentation deployed at present in the routine 

monitoring networks for determination of gas-phase 

NO2 and SO2 concentrations is likely adequate for 

determining compliance with the current NAAQS. But 

in application for determining environmental effects, 

all these methods have important limitations, which 

make them inadequate for fully characterizing the state 

of the atmosphere at present, correctly representing the 

complex heterogeneity of N and S deposition across 

the landscape, and for realistically apportioning 

the contributions of reduced and oxidized forms of 

atmospheric N and S in driving observed biological 

effects at a national scale.

Although the current standard for NO2 is inadequate to 

protect welfare, there is a straightforward technical fix to 

this problem. NOx (NO + NO2) is a variable, often small 

component of reactive oxidized nitrogen (NOy) as has 

been noted in EPA’s Integrated Science Assessment for 

Oxides of Nitrogen and Sulfur – Ecological Criteria and 

extensively documented in the reviewed literature (Fahey 

et al., 1986; Fehsenfeld et al., 1987; Doddridge et al., 1991; 

Parrish et al., 1993; Poulida et al., 1994; Ridley et al., 

1994; Emmons et al., 1997; Liang et al., 1998; Munger et 

al., 1998; Zhou et al., 2002; Takegawa et al., 2003; Parrish 

et al., 2004a,b; Horii et al., 2004, 2006; Dunlea et al., 

2007; Kleinman et al., 2007; Hargrove and Zhang, 2008; 

Luria et al., 2008; Zhang et al., 2008; Schwab et al., 2009; 

Luke et al., 2010). The standard chemiluminescence NO/

NOx technique measures NO directly via reaction with 

O3, and NOx (NO + NO2) by conversion of NO2 to NO

on hot molybdenum. It had been thought that NOx would 

make up nearly all of NOy in heavily polluted urban areas, 

but results including some from Mexico City (Dunlea et 

al., 2007) indicate that this is not the case. Interferences 

were up to 50% of the ambient NO2 concentration even 

in a heavily urban area. Recent results from Houston, TX 

(Luke et al., 2010) showed a median NOx/NOy ratio of 

~0.63 between 1300 and 1500 local time, a time of rapid 

ozone production. NO2 has been used historically as the 

indicator for many health studies, and it is regulated in part 

because of studies that directly link it to respiratory effects. 

However, given the inadequacy of NO2 as a standard to 

protect welfare, the Committee recommends that EPA 

reexamine the criteria pollutant oxides of nitrogen and the 

indicator species NO2 and consider adding chemically 

reactive nitrogen as a criteria pollutant, and NHx and NOy

as indicators to supplement the NO2 National Ambient Air 

Quality Standard. 

The references listed above also document that current 

suite of NOx monitors are useful for proving compliance 

with the current NO2 standard, but suffer substantial 

interferences and the extent of these interferences varies 

with time and location. Reactive N compounds, especially 

HNO 3, can be lost on the inlet components. Because the 

inlet does not transmit all reactive nitrogen compounds 

with high efficiency, commercial NOx monitors may 

provide an upper limit for NOx and a lower limit for 

NOy, but they measure neither NOx nor NOy precisely. 

Numerical simulations do not currently produce consistent 

results for the partitioning of NOy species, for example 

see Archibald et al., (2010). The data from the current 

monitoring network are of limited value for determining 

exposure to NO2, for evaluating chemical transport models 

(CTMs), or for assessing efficacy of emissions control 

strategies (McClenny et al., 2002). 

NOy monitors differ from the NOx monitors currently 

in use in the position of the hot molybdenum NO2 to NO

converter and in calibration and operation (e.g., Thermo 

Scientific Model 42i-Y). In an NOx monitor the converter 

is well downstream of the inlet, behind a filter that 

removes particulate matter, while in an NOy monitor the 

converter is at the inlet (Fehsenfeld et al., 1987; Doddridge 

et al., 1991; Doddridge et al., 1992; Parrish et al., 1993; 

Poulida et al., 1994; Ridley et al., 1994; Emmons et al., 

1997; Liang et al., 1998; Munger et al., 1998; Parrish et 

al., 2004b; Dunlea et al., 2007; Hargrove and Zhang, 2008; 

Luria et al., 2008; Schwab et al., 2009). Thus the current 

monitors could be replaced or retrofitted to measure NOy

with a detection limit of 0.1 ppb at reasonable expense 

for equipment and for training operators. NOy monitors, 

like NOx monitors, provide proof of compliance with 

NO2 standards. For specific measurement of NO2, 

commercial instruments for selective NO2 reduction (e.g., 

Air Quality Design, Inc., Wheat Ridge, CO) or direct 

NO2 measurement are beginning to be available (Parrish 

and Fehsenfeld, 2000; Castellanos et al., 2009) and these 

may offer an alternative approach. Direct monitoring 

of NO2 and other Nr species is a long-term goal, but 

new techniques will require thorough testing in polluted 

environments. NOy monitors using hot molybdenum 

compare favorably with other NOy techniques such as 

gold-catalyzed CO reduction and several studies have 

concluded that NOy can be reliably measured in suburban 

and urban environments (Fehsenfeld et al., 1987; Crosley, 

1996; Williams et al., 1998).

Conclusions on atmospheric deposition of Nr

As discussed here and in Appendix A, downward 

transport from the atmosphere is a major source of 

Nr to the earth’s surface, but there are uncertainties in 
33

the characteristics and absolute magnitude of the flux. 

Pollutants not deposited are exported from the continent 

and alter the composition and radiative balance of the 

atmosphere on a large scale. A review of the literature 

revealed the following major points concerning the 

present state of the science:

1. Measurements from the National Atmospheric 

Deposition Program (NADP) indicate that wet 

deposition of ammonium plus nitrate for the period 

2000-2006 averaged 3.1 kg N/ha/yr over the 48 

contiguous states. 

2. The reduced (NH4

+) and oxidized (NO3-

) forms of 

reactive N contributed about equally to the flux, but 

input to the eastern United States was greater (and less 

uncertain) than to the western U.S. 

3. For the U.S. east of the Mississippi River, dry 

deposition data have also been analyzed – the Clean Air 

Standards and Trends Network (CASTNET) monitors 

vapor phase HNO3, as well as particulate NO3

-

 and 

NH4

+. These measurements indicate 7.75 kg N/ha/

yr total deposition (5.46 wet 2.29 dry) over the East. 

Conspicuous by its absence from this number is dry 

deposition of ammonia. 

4. Decreases in NOx emissions appear to have led to 

decreases in NO3

-

 deposition. NADP data show a 

national decreasing trend in the wet nitrate deposition 

and some individual sites show statistically significant 

decreases in deposition and correlations with emissions. 

5. A thorough review of all published studies of the U.S. 

NOy budget indicates that about 70 % of the NOx

emitted by the U.S. is deposited onto the continent 

with the remainder exported, although substantial 

uncertainty remains. Major sources of uncertainty 

concerning the deposition of NOy include dry 

deposition of unmonitored members of the NOy

family, uncertainties in the chemistry of organic N, and 

poorly constrained estimates of convective venting of 

the planetary boundary layer of the atmosphere (i.e., 

uncertainties concerning the venting of NOy from the 

lowest layer of the atmosphere). 

6. Based on observations and model estimates of the 

relative deposition of unmeasured quantities, total 

estimated deposition of all forms of Nr for the period 

2000-2004 is ~11 kg N /ha /yr for the eastern U.S., and 

for the 48 states ~7.5 kg N /ha /yr with a range of 5.5 to 

9.5 kg N /ha /yr. 

Finding 8: Scientific uncertainty about the origins, 

transport, chemistry, sinks, and export of Nr remains high, 

but evidence is strong that atmospheric deposition of Nr to 

the earth’s surface as well as emissions from the surface to 

the atmosphere contribute substantially to environmental 

and health problems. Nitrogen dioxide, NO2, is often a 

small component of NOy, the total of oxidized nitrogen 

in the atmosphere. The current NAAQS for NO2, as an 

indicator of the criteria pollutant “oxides of nitrogen,” 

is inadequate to protect health and welfare. NOy should 

be considered seriously as a supplement or replacement 

for the NO2 standard and in monitoring. Atmospheric 

emissions and concentrations of Nr from agricultural 

practices (primarily in the form of NH3) have not been well 

monitored, but NH4

+ ion concentration and wet deposition 

(as determined by NADP and NTN ) appear to be 

increasing, suggesting that NH3 emissions are increasing. 

Both wet and dry deposition contribute substantially to 

NHx removal from the atmosphere, but only wet deposition 

is known with much scientific certainty. Thus consideration 

should be given to adding these chemically reduced and 

organic forms of Nr to the list of Criteria Pollutants. 

Recommendation 8a: EPA should reexamine the 

criteria pollutant “oxides of nitrogen” and the indicator 

species NO2 and consider adding chemically reactive 

nitrogen as a criteria pollutant, and NHx and NOy as 

indicators to supplement the NO2 National Ambient Air 

Quality Standard.

Recommendation 8b: Monitoring of NHx and NOy

should begin as soon as possible to supplement the 

existing network of NO2 compliance monitors. 

Recommendation 8c: EPA should pursue the longerterm goal of monitoring individual components of Nr, such 

as NO2 (with specificity), NO and PAN, and HNO3, and 

other inorganic and reduced forms, as well as support the 

development of new measurement and monitoring methods. 

Recommendation 8d: The scope and spatial coverage 

of the Nr concentration and flux monitoring networks 

(such as the National Atmospheric Deposition Program 

and the Clean Air Status and Trends Network) should be 

increased and an oversight review panel for these two 

networks should be appointed.

Recommendation 8e: EPA in coordination with other 

federal agencies should pursue research goals including:

• Measurements of deposition directly both at the 

CASTNET sites and in nearby locations with nonuniform surfaces such as forest edges.

• Improved measurements and models of convective 

venting of the planetary boundary layer and of long 

range transport. 

• Improved analytical techniques and observations 

of atmospheric organic N compounds in vapor, 

particulate, and aqueous phases.

• Increased quality and spatial coverage of 

measurements of the NH3 flux to the atmosphere 

from major sources especially agricultural practices.

• Improved measurement techniques for, and numerical 

models of NOy and NHx species (especially with 

regard to chemical transformations, surface 

deposition and offshore export, and linked oceanland-atmosphere models of Nr).
34

Input and recycling of Nr within terrestrial 

systems in the United States

Annual input of newly created Nr onto terrestrial 

ecosystems comes primarily from atmospheric 

deposition, synthetic fertilizer and BNF in managed 

and unmanaged ecosystems (Table 1). Although Nr 

deposited from the atmosphere to terrestrial systems is 

formed inadvertently during fossil fuel combustion and 

from volatilization of NH3 from agricultural activities, 

it serves to provide nutrients, along with biological 

N fixation and synthetic fertilizer, for food, feed, and 

fiber production in the agricultural sector. Forests and 

grasslands use Nr for growth. Home gardens, parks and 

recreational areas utilize Nr within the urban landscape. 

Approximately 32 Tg of new Nr reached the land of the 

48 contiguous states in 2002 (Table 1). An additional 

~0.2 Tg of N was imported mainly as food and drink 

products (FAO, 2007). An additional ~12 Tg of Nr 

was recycled back to terrestrial and aquatic systems in 

livestock excreta (~6 Tg N), human excreta (~2 Tg N), 

and crop residue from the previous year’s production (~4 

Tg N; U.S. EPA, 2007e). Of this N, ~ 1.3 Tg (~1.2 from 

livestock manure and <0.1 from sewage sludge) was 

used as fertilizer for crop production (U.S. EPA, 2007e). 

More detailed information describing sources and 

cycling of reactive nitrogen input in terrestrial systems is 

provided in Appendix B. 

Finding 9: Total N budgets within all terrestrial 

systems are highly uncertain. The relative amount of 

N losses from terrestrial systems ascribed to leaching, 

runoff, and denitrification are as uncertain as the N

budgets themselves.

Recommendation 9: EPA should join with USDA, 

DOE, and universities to work together in efforts to 

ensure that the N budgets of terrestrial systems are 

properly quantified and that the magnitudes of at least the 

major loss vectors are known.

Transfer of Nr to aquatic systems

Within the nitrogen cascade, Nr flows from the 

atmosphere and terrestrial systems into aquatic systems. 

Aquatic systems include groundwater, wetlands, 

streams and rivers, lakes and the coastal marine 

environment. Nr is deposited directly into surface 

aquatic systems from the atmosphere (direct deposition) 

and Nr that is not either stored or removed as products 

on terrestrial systems eventually moves into aquatic 

systems (indirect deposition). 

The area of an airshed generally greatly exceeds that 

of a watershed for a specific estuary or coastal region. 

For example, the airshed of the Baltic Sea includes 

much of western and central Europe (Asman, 1994; 

Hov et al., 1994), while the airsheds of the two largest 

U.S. estuarine ecosystems, the Chesapeake Bay and 

Albemarle-Pamlico Sound, are 15 to over 30 times the 

size of their watersheds (Dennis, 1997). Thus, the airshed 

of one region may impact the watershed and receiving 

waters of another, making eutrophication a regional-scale 

management issue (Galloway and Cowling, 2002; Paerl et 

al., 2002). Furthermore, atmospheric N inputs do not stop 

at coastal margins. Along the North American Atlantic 

continental shelf, atmospheric N inputs more than match 

riverine inputs (Jaworski et al., 1997; Paerl et al., 2002), 

underscoring the fact that N-driven marine eutrophication 

may require regional or even global solutions. Even in 

truly oceanic locations (e.g., Bermuda), North American 

continental atmospheric N emissions (reduced and 

oxidized N) are commonly detected and significant (Luke 

and Dickerson, 1987; Prospero et al., 1996). Likewise, 

islands in the North Pacific receive N deposition 

originating in Asia (Prospero et al., 1989).

Riverine and atmospheric “new” Nr inputs in the 

North Atlantic Ocean basin are at least equal and may 

exceed “new” Nr inputs from biological N2 fixation 

(Howarth et al. 1996; Paerl and Whitall, 1999; Paerl et 

al. 2002). Duce et al. (2008) estimate that up to onethird of the oceans’ external Nr supply enters through 

atmospheric deposition. This deposition leads to an 

estimated ~ 3% of new marine biological production and 

increased oceanic N2O production. Schlesinger (2009) 

estimated that global atmospheric transport of Nr from 

land to sea accounts for the movement of almost one 

third of the annual terrestrial Nr formation. Therefore, 

our understanding of marine eutrophication dynamics, 

and their management, needs to consider a range of scales 

reflecting these inputs, including ecosystem, watershed, 

regional and global levels.

One example of shifting N inputs is the proliferation 

of intensive livestock operations in coastal watersheds, 

which has led to large increases and changes in chemical 

composition of nitrogenous compounds discharged to 

estuarine and coastal waters via runoff, groundwater, and 

atmospheric deposition (Paerl, 1997; Howarth, 1998; 

Galloway and Cowling, 2002). In general, coastal waters 

under the influence of these operations are experiencing 

increases in total N loading as well as a shift toward more 

reduced N (NH4

+, organic N) relative to oxidized N (NO3-

) (Galloway and Cowling, 2002; Howarth et al., 2002). 

These increases, combined with increases in hypoxia and 

anoxia in receiving waters, are leading to more NH4

+-

rich conditions, which will favor those algal groups best 

able to exploit this N form, including some harmful 

algal bloom (HAB) taxa (Paerl and Whitall, 1999; 

Paerl et al., 2007). Similarly, conversion of forest and 

agricultural lands to urban lands can alter landscapes and 

promote N loading to estuaries by increasing impervious 

pathways and removing natural landscape filters for Nr. 

Development also destroys wetlands, leading to more 

NO3

-

-enriched conditions, potentially favoring plant taxa 

best able to exploit this N form. 

A recent evaluation of decadal-scale changes of 

NO3

-

 concentrations in ground water supplies indicates 
35

that there has been a significant increase in nitrate 

concentrations in well water across the U.S. (Rupert, 

2008). This study compared the nitrate content of 495 

wells during 1988-1995 with nitrate content found during 

2000-2004 as a part of the U.S. Geological Survey, 

National Water Quality Assessment Program. In a subset 

of wells, ground water recharge was correlated with 

historic fertilizer use and it was concluded that nitrate 

concentrations in ground water increased in response to 

the increase of N fertilizer use.

2.3.2. Storage of Nr within Terrestrial 

Environmental Systems 

According to the nitrogen cascade conceptualization, 

terrestrial environmental systems are compartmentalized 

into agriculture, populated, and vegetated systems. 

Annual input of Nr is greatest in agricultural ecosystems 

(farmland, cropland, and grazed pastureland). Annual Nr 

inputs to agricultural ecosystems, using 2002 as the base 

year, include 9.8 Tg from synthetic fertilizer, 7.7 Tg from 

biological N fixation in crops (mainly soybeans), and 1.3 

Tg from atmospheric deposition. Nr input into vegetated 

systems (mostly forested, but including non-cropland 

grasslands and other natural vegetation types as well) comes 

mostly from atmospheric deposition (3.2 Tg). Annual input 

of Nr into populated systems includes synthetic fertilizer 

application to urban turfgrass and recreational areas (~1.1 

Tg), and atmospheric deposition 0.2 Tg.

Much of the annual Nr input into these terrestrial 

systems passes through, and is transferred within, 

terrestrial systems or atmosphere via NH3, NOx or N2O, 

or aquatic environmental systems via NO3

-

 and organic N

leaching and runoff or NHx and NOy deposition.

The largest single reservoir of total N in the terrestrial 

environmental system is soil organic matter (SOM). 

Approximately 52,000 Tg C and 4,300 Tg N are contained 

in the upper 100 cm of soil in the 48 contiguous states (N

is estimated from assumed C/N ratio of 12) (Lal et al., 

1998). For comparison, the total above ground biomass 

of U.S. forests of these states contains ~ 15,300 Tg of C

and ~ 59 Tg N (estimated using a C/N ratio of 261), and 

15,500 Tg of soil organic matter carbon (SOM-C), and 

1,290 Tg total N (estimated using a C/N ratio of 12) (U.S. 

EPA, 2007e). Most of this soil organic matter nitrogen 

(SOM-N) is bound within complex organic molecules 

that remain in the soil for tens to thousands of years. A 

small fraction of this SOM is mineralized, converted to 

carbon dioxide and Nr annually. The total N contained 

within above and below ground compartments isn’t really 

of concern. What is of interest in addressing issues of 

Nr, is the change in N stored within the compartments 

of terrestrial systems. The pertinent question is whether 

N is being retained or released from long-term storage. 

The Committee evaluated estimates of annual change 

of N storage within important components of terrestrial 

systems. The Committee used Carbon stock information 

in Chapter 7 of the EPA Inventory of U.S. Greenhouse 

Gas Emissions and Sinks 1990-2005 (U.S. EPA, 2007e) 

to estimate N storage in U.S. terrestrial systems. Nitrogen 

stock change was determined by simply assigning a molar 

C/N ratio of 12 for soils and 261 for trees and making the 

appropriate conversions from C to N.

Box 1: Hypoxia in the Gulf of Mexico

An example of a problem of excess Nr that moves from one part of the U.S. to another is the movement of Nr 

from the states that make up the Mississippi River drainage to the Gulf of Mexico (see discussion in Section 5.3.4). 

A hypoxic zone covers a significant area of the receiving bottom waters of the continental shelf of the northern 

Gulf of Mexico (details may be gleaned from U.S. EPA SAB, 2007). This is a seasonally severe problem that has 

persisted there for at least the past 20 years. Between 1993 and 1999 the hypoxia zone ranged in extent from 13,000 

to 20,000 km2 (Rabalais et al. 1996, 1999; Rabalais and Turner, 2001). The hypoxia is most widespread, persistent, 

and severe in June, July, and August, although its extent and timing can vary, in part because of the amplitude and 

timing of flow and subsequent nutrient loading from the Mississippi River Basin. The waters that discharge to the 

Gulf of Mexico originate in the watersheds of the Mississippi, Ohio, and Missouri Rivers (collectively described 

here as the Mississippi River Basin). With a total watershed of 3 million km2, this basin encompasses about 40% of 

the territory of the lower 48 states and accounts for 90% of the freshwater inflow to the Gulf of Mexico (Rabalais et 

al. 1996; Mitsch et al., 2001; U.S. EPA SAB, 2007). 

The December, 2007 report, Hypoxia in the Northern Gulf of Mexico: An update by the EPA Science Advisory 

Board (U.S. EPA SAB, 2007) determined that, “To reduce the size of the hypoxic zone and improve water quality 

in the Basin, the SAB Panel recommends a dual nutrient strategy targeting at least a 45% reduction in riverine total 

nitrogen flux (to approximately 870,000 metric tons/yr) and at least a 45% reduction in riverine total phosphorus 

flux (to approximately 75,000 metric tons/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.”
36

Agricultural systems

Croplands within the contiguous 48 states occupy 

~ 368 million acres (149 million ha) (19%) of the 785 

million ha of total land area. In 2002, 126 million ha 

of this cropland were cultivated (USDA NRCS, 2007). 

Croplands are generally found on well drained mineral 

soils (organic C content 1-6% in the top 30 cm). Small 

areas of drained organic soils are cultivated (organic C

content of 10-20%) in mainly Florida, Michigan and 

Minnesota (EPA, 2007e). Organic soils lost ~0.69 Tg of 

Nr in 2002 while mineral soils accumulated ~1.5 Tg of Nr 

(Table 10). Much of the accumulation of SOM-C was due 

to the use of conservation tillage and high yielding crop 

varieties (U.S. EPA, 2007e). Losses of Nr from organic 

soils are due to mineralization of SOM and release of 

Nr input. In cultivated soils annual input of new Nr is 

approximately 9.7 Tg from fertilizer N, 1.1 Tg from 

livestock manure (recycled N), ~7.7 Tg from biological 

N fixation, and 1.2 Tg from atmospheric deposition. 

Assuming that loss of fertilizer N from the small area of 

organic soils is a minor fraction of the total, ~17% of N

input from synthetic fertilizer, ~12% of total N input, is 

stored in cropland mineral soils annually. 

According to the U.S. EPA National Greenhouse 

Inventory (U.S. EPA, 2007g) the net increase in soil 

C stocks over the period from 1990 through 2005 was 

largely due to an increase in annual cropland enrolled in 

the Conservation Reserve Program, intensification of crop 

production by limiting the use of bare-summer fallow in 

semi-arid regions, increased hay production, and adoption 

of conservation tillage (i.e., reduced- and no-till practices). 

It is clear that conversion of marginal crop land to CRP

stores C and reduces erosion and nitrate leaching. Likewise, 

use of soil conservation tillage, as opposed to conventional 

tillage with a plow or disk, reduces erosion. However, 

the impact of conservation tillage on soil C storage and 

Nr losses due to leaching or denitrification are much less 

certain. For example, although the EPA estimates shown 

in Table 10 assume that no-till crop production results 

in net carbon sequestration, recent publications indicate 

that no-till cropping practices do not result in net carbon 

sequestration (Verma et al., 2005; Baker et al. 2007; 

Blanco-Canqui and Lal, 2008). Therefore, the estimates of 

soil C and N storage in mineral soils in Table 10 that were 

derived from U.S.EPA, (2007b) need to be reconsidered. 

These new studies and that of David et al. (2009) suggest 

that organic C conservation by reduced tillage practices has 

been overestimated because soil sampling and analysis has 

been confined to the top 30 cm of soil when the top meter 

of soil needs to be considered. Baker et al. (2007) and 

Verma et al. (2005) also show that long-term, continuous 

gas exchange measurements have not detected C gain due 

to no-till practices. They concluded that although there are 

other good reasons to use no-till, evidence that it promotes 

C sequestration is not compelling. These findings highlight 

the need for appropriate assessment of ecosystem N storage 

in order to confirm or disprove this Committee’s conclusion 

that only a small part of annual Nr input is stored in 

agricultural lands, forests, and grasslands.

Populated systems – urban lands

Populated or “developed land” (developed land is the 

terminology used by the U.S Department of Agriculture’s 

Natural Resources Conservation Service [NRCS]) occupied 

~106 million acres (42.9 million ha) of the U.S. land area 

in 2002. This equates to approximately 5.5% of the U.S. 

land area (USDA NRCS, 2007). The EPA Inventory of U.S. 

Greenhouse Gas Emissions and Sinks (U.S. Environmental 

Protection Agency, 2007e) indicates that urban areas cover 

over 4.4% of the land area with tree canopy covering 

27.1% of the urban area. The tree-covered area constitutes 

approximately 3% of total tree cover in the continental U.S. 

If the NRCS value of 42.9 million ha is used, then trees 

cover ~27.9 million acres (11.3 million ha) of urban land in 

the contiguous 48 states. Another ~ 35.1 million acres (14.2 

million ha) of land is covered by turf grass in parks, golf 

courses, and lawns. In both urban forests and turf grass, Nr 

storage is dependent upon the age of the trees or turf. In 

young (pre-steady state) systems N is being accumulated, 

while at steady state no net change occurs. Some areas may 

be degrading and actually losing biomass and returning N

to the environment. EPA does not estimate carbon changes 

in turf grass, but does estimate changes in carbon storage in 

urban forests (U.S. EPA, 2007e). Urban trees sequestered 

an estimated net 22 Tg of carbon and 0.12 Tg of N in 2002 

(assuming that urban trees are mainly hardwood having 

a C/N ratio of 186) (U.S. EPA, 2007e). Annual fertilizer 

N input into the urban landscape is approximately 10% 

of total fertilizer N consumption in the U.S. (U.S. EPA, 

2007e), or ~1 Tg of N in 2002. Another 0.2-1.0 Tg N is 

deposited from the atmosphere. In some areas deposition 

can be disproportionately high due to locally high NOy

concentrations. Storage of ~0.12 Tg N in urban forests 

constituted approximately 3% of Nr input annually.

Vegetated systems – forests and grasslands

Forests cover approximately 164 million ha, ~21% 

of the land area of the contiguous 48 states (USDA 

NRCS, 2007). The forest carbon stocks analysis by 

EPA (U.S EPA, 2007e) is based on state surveys that 

are conducted every 1 to 10 years. Annual averages 

are applied to years between surveys. To determine 

changes in C related to the rate of tree growth, Birdsey 

(1992) estimated that there were 52,500 Tg of C above 

and below ground in U.S. forests; soil contained 59% 

of total C, 9% was in litter, and 5% in tree roots. The 

EPA estimate for 2002 was 43,600 Tg of C. To estimate 

N storage based on EPA data (U.S. EPA, 2007e), the 

Committee has assumed that forests are 85% softwood 

and 15% hardwood with an average C/N ratio of 261. 

These estimates indicate that forests and forest products 

stored ~0.43 Tg of N in 2002 (Table 10).
37

Grasslands, including rangelands and pasturelands, 

occupy approximately 213 million ha (27.1%) of the 

contiguous 48-state land area. The NRCS divides 

these grasslands into pastureland (48.2 million ha) and 

rangeland (164 million ha). Pastureland is managed (it 

may be fertilized and mown) and rangeland is managed 

only to the extent that livestock grazing intensity is 

regulated on the land used for such grazing. Changes in 

the N status of grasslands are dependent upon changes in 

soil organic matter as the above ground biomass produced 

annually is either consumed by livestock or decomposed 

in the field. Soil organic C stocks were estimated using 

the Century biogeochemical model and data used were 

based upon the NRCS/National Resources Inventory 

(NRI) survey (U.S. EPA, 2007e). Changes in soil N

content were estimated using a C/N ratio of 12. Nitrogen 

input into rangelands is generally only from atmospheric 

deposition, which contributes 1.9 Tg N each year to range 

production. Rangeland tends to be in relatively remote 

areas where atmospheric Nr deposition is low.

Collectively, forests and grasslands stored ~0.74 Tg 

of N in 2002. Much of the soil N storage in grasslands is 

a result of conversion of croplands to grasslands, mainly 

due to the conservation reserve program. Forest soils 

Table 10: Net annual change in continental U.S. croplands soil N and C, forest C and N, and 

grasslands C and N in 2002 

 C N

Cropland

Cropland remaining cropland

Mineral soil 17 1.4*

Organic soil -8.3 -0.69

Land converted to cropland 0.8 0.067

Total 9.5 0.80

Forests

Forests and harvested wood products

Above ground biomass 85 0.32

Below ground biomass 16 0.063

Dead wood 9.1 0.035

Litter 7.2 0.028

Soil organic matter -2.8 -0.23

Harvested Wood 59 0.22

Total 174 0.44

Grasslands

Grasslands remaining grasslands

Mineral soil -0.8 -0.067

Organic soil -1.3 -0.11

Lands Converted to Grasslands 5.8 0.48

Total 3.7 0.31

US Total C & N Storage in 2002 187 1.5

Measurements in Tg. Negative sign indicates a decrease in storage; positive number indicates increase in storage. Soil 

C/N ratio = 12; wood C/N = 261. C storage numbers were obtained from U.S. EPA, 2007e.* See previous discussion of 

soil organic matter accumulation in croplands.
38

appear to be losing N while overall N storage in forests is 

from accumulation in above-ground biomass and that that 

remains in forest products that are stored for long periods.

Summary of estimates of Nr stored in terrestrial 

systems in 2002

Although our estimate of N storage in terrestrial 

systems is highly uncertain, annual N storage is likely 

relatively small. Approximately 1.5 Tg N was stored in 

terrestrial systems of the contiguous 48 states in 2001 

(Table 10). Soils were the largest reservoir with croplands 

(0.8) and grasslands (0.3) sequestering most of the N. 

Annual storage in agricultural, grassland and forest 

soil and in forest biomass is approximately 6 to 10% of 

annual Nr input. Estimated total Nr input from synthetic 

fertilizer, biological N fixation, and atmospheric deposition 

into terrestrial systems within the contiguous 48 states in 

2002 was ~32 Tg. All of the input and outflow numbers 

are highly uncertain, but N loss through denitrification 

appears to be the major loss mechanism. Storage in soils 

and trees appears to account for only a small portion of the 

annual N input while apparent loss through denitrification 

dominates the budget. This is discussed in Section 2.3.3 of 

this report and in a recent global Nr review by Schlesinger 

(2009). Some small fraction is re-volatilized and exported 

from the continent. 

2.3.3. Areas of Uncertainty in Nr Transfer 

and Transformation

In considering Nr transfers and transformations in 

and between the environmental systems of the nitrogen 

cascade, the Committee has encountered a number of 

areas where quantities or flows of Nr are highly uncertain. 

All of these areas need attention from EPA in conjunction 

with other federal and state agencies and universities. 

Although most of the following points have been 

highlighted in various findings and recommendations 

within other chapters of this report, we feel the need to 

highlight the following areas:

n Total denitrification in animal feeding operations, in 

soils, and in aquatic systems needs to be quantified 

along with all gaseous products produced and released 

to the atmosphere during nitrification/denitrification. 

These gases include NOx, N2O and N2.

n The amount of Nr transferred to each environmental 

system as dry deposition needs to be quantified and 

monitored.

n The fraction of NOy in the form of organo-nitrates and 

other organic nitrogen species is poorly quantified, but 

may play a major role in air quality and Nr cycling.

n Rates and amounts of ammonia emissions from 

fertilized soils and animal feeding operations need to be 

quantified and the fate of this ammonia determined.

n The annual change in N storage in soils (agricultural, 

forest, grassland and urban areas) needs to be quantified 

in conjunction with the change in carbon.

These areas of high uncertainty are highlighted 

because very little information exists in some of the areas. 

In other areas, such as denitrification and the relative 

release of N2O from soils and aquatic systems, the sparse 

data are highly variable and this makes developing 

meaningful guidelines for control difficult. An analysis of 

Nr input and fate in 16 watersheds in the U.S. is provided 

as an example to illustrate how the inputs and fate of Nr 

can be evaluated for a large watershed.

Input and fate of Nr in 16 watersheds in the 

northeast United States

There are no comprehensive data available to assess 

the transfer and transformations in and between the 

atmosphere, terrestrial systems (agriculture, populated, 

and vegetated systems) and aquatic systems nationally. 

Determining a national N budget is a priority research 

area. As there are no national data available, an example 

analysis of Nr input and fate in 16 watersheds in the 

northeast U.S. (for which data are available) is used to 

show an evaluation of the inputs and fate of Nr for a large 

watershed (Van Breemen et al., 2002). 

The watersheds in this study encompass a range of 

climatic variability, Maine to Virginia. The watersheds are 

a major drainage to the coast of the North Atlantic Ocean. 

Using data from the early 1990s, Boyer et al. (2002) 

reported the quantification of N inputs to each watershed 

from atmospheric deposition, N fertilizers, biological N

fixation, and import of N in agricultural products (food 

and feed). They compared inputs with N losses from the 

system in riverine export. As a part of the same study, 

Van Breemen et al. (2002) analyzed the fate of N inputs 

to these watersheds and developed budgets for each 

watershed. The total area of the watersheds was 32,666 

square km with land use categories of forest (72%), 

agricultural (19%), urban (3%), wetlands (5%), and 

1% other uses. The Nr input into the watersheds (using 

weighted averages for all 16 watersheds) was 3,420 kg 

per square km per year. Figure 14 shows the Nr sources 

and the estimated fate of this Nr as a per cent of the 

weighted average Nr input.

Van Breemen et al. (2002) indicate that Nr inputs and 

storages and losses were well correlated (R2 = 0.98). 

Denitrification in landscape soils is the most uncertain 

estimate because rates are calculated by difference 

between total inputs and outputs so they accumulate 

errors from all estimates. The denitrification loss term 

may also reflect the change in N storage in groundwater. 

The net storage of N in the soil (18% of total storage and 

losses) indicates that there is a non-steady state condition 

in the soil. Increasing storage of Nr on land implies that 

drainage and denitrification exports of Nr are likely to 

increase when a new steady-state condition is reached.

These data suggest that Nr research needs to focus on 

understanding the “denitrification” loss term in this analysis. 

The losses occur in the terrestrial landscape, before Nr 
39

enters the river. Where do these losses occur, within the 

agricultural field, in drains and ditches near the agricultural 

field, in riparian areas, or wetlands? Understanding this 

term may help in the management of Nr in watersheds to 

decrease nitrate movement into aquatic systems as well as to 

limit N2O emissions to the atmosphere.

The Van Breemen et al. (2002) study also estimated 

that approximately 30% of N input was exported to the 

rivers and about two thirds (20% of total N input) of this 

N was exported to coastal waters by rivers. The remaining 

one-third (11% of total N input) was considered to 

have been denitrified in the rivers. These examples 

also demonstrate that Nr in the atmosphere, terrestrial 

systems and aquatic systems are not separate and must 

be considered collectively. Atmospheric deposition is 

a variable but important input into aquatic systems that 

contributes to Nr enrichment problems. Aquatic and 

terrestrial systems process this Nr and return other Nr 

gases (NH3, NOx and N2O to the atmosphere). Nr from 

terrestrial systems impacts both the atmosphere and 

aquatic systems through emission of NH3, NOx, N2O and 

leaching and runoff of NO3

-

.

Finding 10: Denitrification of Nr in terrestrial and 

aquatic systems is one of the most uncertain parts of the 

nitrogen cycle. Denitrification is generally considered 

to be a dominant N loss pathway in both terrestrial and 

aquatic systems, but it is poorly quantified.

Recommendation 10: EPA, USDA, DOE, and 

universities should work together to ensure that 

denitrification in soils and aquatic systems is properly 

quantified, by funding appropriate research.

Atmospheric

Deposition

33%

Fertilizer

15%

Forest Fixation

5%

Net Import in

Food & Feed

24%

Agricultural Fixation

23%

Denitrification

in Landscape

37%

Ammonia

Volatilization

3% Dentrification

in Rivers

11%

Riverine Export

20%

Wood Export

Soil Storage 5%

Increase

9%

Food

Export

6%

Biomass

Increase

9%

Nr Sources

Nr Storage & Loss

Figure 14: Nr input and loss from 16 watersheds 

in the northeast United States 

Source: Van Breeman et al., 2002 (Figure 7, p. 289). 

Reprinted with permission; copyright 2002, Springer 

Science+Business Media B.V.
40
41

This chapter summarizes knowledge of the impacts 

of Nr on freshwater, coastal, atmospheric, and terrestrial 

ecosystems. Table ES-1 in the Executive Summary 

provides a more detailed presentation of the quantitative 

extent of Nr impacts on the environment.

3.1. Impacts on Drinking Water, Human 

Health, and Freshwater Biota

A detailed presentation on the impacts of Nr on 

various aquatic systems is presented in Appendix G. The 

EPA’s Office of Water (U.S. EPA, 2007b) has noted the 

following impacts caused by excessive Nr in aquatic 

systems:

n Excessive nutrients (N and P) can cause negative 

ecological impacts to water bodies on a national scale 

by stimulating harmful algal blooms.

• Algal blooms block sunlight and result in the 

destruction of submerged aquatic vegetation which 

serves as critically important habitat and food for 

many organisms.

• Algal blooms eventually die off and consume 

dissolved oxygen from the water column which can 

lead to die off of aquatic organisms.

• One result of algal blooms is decreased biological 

diversity and populations, including smaller 

populations of game and commercial fish.

• Some blooms, considered harmful algal blooms or 

HABs, have a toxic effect on living organisms and 

are disruptive of ecosystem structure and transfer of 

energy to higher trophic levels.

n Excessive nutrients also pose public health risks.

• Algal blooms can cause taste and odor problems in 

drinking water.

• Hazardous algal blooms can cause respiratory distress 

and neurological problems in swimmers.

• Excessive nitrates can cause “blue baby syndrome.”

n Nutrient pollution is occurring at a national scale and 

has not been completely addressed.

• 49 states and 4 territories have 303(d) listings due to 

nutrients, and about 50% of the states have greater 

than 100 water quality impairments due to nutrients.

• Over 10,000 impairments are a result of nutrient 

pollution.

Finding 11: There is growing recognition of 

eutrophication as a serious problem in aquatic systems 

(NRC, 2000). The last comprehensive National Coastal 

Condition Report was published in 2004 (U.S. EPA, 

2004) and included an overall rating of “fair” for 

estuaries, including the Great Lakes, based on evaluation 

of more than 2,000 sites. The water quality index, which 

incorporates nutrient effects primarily as chlorophyll-a 

and dissolved oxygen impacts, was also rated “fair” 

nationally. Forty percent of the sites were rated “good” 

for overall water quality, while 11% were “poor” and 

49% “fair.” 

Recommendation 11. The Committee recommends 

that EPA develop a uniform assessment and management 

framework that considers the effects of Nr loading 

over a range of scales reflecting ecosystem, watershed, 

and regional levels. The framework should include all 

inputs related to atmospheric and riverine delivery 

of Nr to estuaries, their comprehensive effects on 

marine eutrophication dynamics and their potential for 

management.

Finding 12: Meeting Nr management goals for 

estuaries, when a balance should be struck between 

economic, societal, and environmental needs, seems 

unlikely under current federal law. Enforceable authorities 

over nonpoint source, stormwater, air (in terms of critical 

loads), and land use are not adequate to support necessary 

Nr controls. Funding programs are presently inadequate to 

meet existing pollution control needs. Furthermore, new 

technologies and management approaches are required 

to meet ambitious Nr control needs aimed at restoring 

national water quality. 

Recommendation 12. The Committee recommends that 

EPA reevaluate water quality management approaches, 

tools, and authorities to ensure Nr management goals 

are attainable, enforceable, and the most cost-effective 

available. Monitoring and research programs should 

be adapted as necessary to ensure they are responsive 

to problem definition and resolution, particularly in 

the development and enhancement of nitrogen removal 

technologies and best management practices, and continue 

to build our level of understanding and increase our ability 

to meet management goals. 

3

Impacts of Nr on Aquatic, Atmospheric, 

and Terrestrial Ecosystems 
42

3.2. Impacts of Airborne or Atmospherically Deposited Nr on Human Health and 

Ecosystems

Six major atmospheric effects are associated with 

increased NOx and NH3 emissions, and two with N2O 

emissions (Galloway et al., 2003). For NOx and NH3

emissions, the effects are:

n Decreases in atmospheric visibility caused by fine PM 

n Elevated ozone concentrations that enhance the 

greenhouse potential of the atmosphere

n Serious ozone and fine particulate matter impacts on 

human health (Pope et al., 1995; Pope, 2000a,b ; Brook 

et al., 2003; Brunekreef et al., 2005; Pope, 2009)

n The important role that NH3 plays in the direct and 

indirect effects of aerosols on radiative forcing and 

thus on global climate change (Penner et al., 1991; 

Seinfeld and Pandis, 1998; Lelieveld et al., 2001; 

Myhre, 2009)

n Decreased productivity of crops, forests, and natural 

ecosystems caused by ozone deposition

n Atmospheric deposition of NH, NH3, NOy, and 

organic forms of Nr that can contribute to ecosystem 

acidification, fertilization, and eutrophication

For N2O, the effects are the greenhouse effect in the 

troposphere and O3 depletion in the stratosphere. 

3.3. Impacts of Nr on Terrestrial 

Ecosystems

As previously discussed, in many terrestrial ecosystems 

the supply of biologically available Nr is a key factor 

controlling the nature and diversity of plant life, and vital 

ecological processes such as plant productivity and the 

cycling of carbon and soil minerals. Human activities 

have not only increased the supply but enhanced the 

global movement of various forms of nitrogen through air 

and water. Appendix H presents information on the impact 

of Nr saturation on ecosystem function.

The primary source of excess Nr for most unmanaged 

terrestrial ecosystems is atmospheric deposition. This 

additional Nr causes a wide variety of sometimes 

beneficial effects (increased growth and productivity of 

forests, natural grasslands, and crops planted in nutrient 

deficient soils) and also sometimes adverse effects on 

terrestrial and aquatic ecosystems in many parts of the 

U.S. Forests and grasslands exposed to excess Nr can 

respond in numerous ways. General effects include the 

following (Woodman and Cowling, 1987; Cowling, 1989; 

Garner et al., 1989; Cowling et al., 1990; Vitousek et al., 

1997a,b; Cowling et al., 2002):

n Increased productivity of forests soils, most of which 

are Nr-limited throughout the U.S. Nr deficiency of 

forest soils has been most fully quantified for pine 

forests in 14 southeastern states.

n Acidification of forest soils leading to decreased 

availability of nutrient cations including calcium, 

magnesium, and potassium and aluminum toxicity, 

established most clearly in the eastern U.S. and both 

central and northern Europe.

n Nr saturation of forest soils (which results in increased 

Nr release to the water draining the soils), presently 

occurring mainly in high-elevation forests of the 

eastern U.S. and southeastern Canada.

n Ozone-induced predisposition of forest trees to damage 

by fungal diseases and insect pests, most clearly 

established in the case of root disease and bark beetles 

in the pine forests of southern California.

n Ozone-induced inhibition of photosynthesis in both 

softwood and hardwood tree species most clearly 

established in controlled exposure studies in both the 

U.S. and Europe at ambient concentrations of ozone 

above 60 ppb. Such concentrations occur frequently 

throughout the eastern U.S. and southeastern Canada.

n Ozone-induced direct injury to foliage, most clearly 

established in the case of “emergence tip burn” in 

eastern white pine.

n Acidification-induced decrease in frost hardiness of 

high-elevation conifer forests, most clearly established 

in the case of red spruce in the northeastern U.S.

n Acidification-induced alteration of beneficial symbiotic 

relationships in forest soils, especially mycorrhizae, 

most clearly established in both northern and central 

Europe.

n Biodiversity losses in natural grasslands and forest 

areas caused by Nr-induced decreases in abundance 

of Nr-limited tree and grass species and replacement 

by Nr-loving weed species, most clearly established in 

both Minnesota and California, and even more vividly 

in The Netherlands.

n Decreases in visibility and increased haziness of the 

atmosphere at scenic vistas in national and state parks 

and wilderness areas.

n More leaching of Nr to aquatic systems via both 

groundwater and surface runoff—a cascade effect.
43

4

Metrics and Implications for Risk Reduction 

Strategies for Reactive Nitrogen

It is important to develop risk reduction strategies 

for reactive nitrogen that take into consideration the 

ways in which Nr is introduced and transformed in 

the environment. This chapter reviews current and 

historical measurement and risk reduction activities 

for Nr and provides specific Committee findings and 

recommendations.

4.1. Measurement of Nr in 

the Environment

Although nitrogen is among the most abundant 

elements on earth, only a small fraction, Nr, is responsible 

for impacts on the environment. Most regulations focus 

narrowly on specific chemical forms of nitrogen as they 

affect media- or site-specific problems, setting limits or 

specifying control technologies without regard to the 

ways in which N is transformed once introduced into 

the environment. Measurement methods are typically 

expressed in terms of mass loadings or concentrations of 

a particular form of N (e.g., ppm NOx, mg/L total NHx, or 

kg/ha of NO3

-

). 

Finding 13: The Committee finds that there is a need 

to measure, compute, and report the total amount of Nr 

present in impacted systems in appropriate units. What 

is measured influences what we are able to perceive and 

respond to; in the case of Nr, it is especially critical to 

measure total amounts and different chemical forms, at 

regular intervals over time. 

Recommendation 13: The Committee recommends 

that EPA routinely and consistently account for the 

presence of Nr in the environment in forms appropriate 

to the medium in which they occur (air, land, and 

water) and that accounting documents be produced and 

published periodically (for example, in a fashion similar 

to National Atmospheric Deposition Program [NADP] 

summary reports). The Committee understands that 

such an undertaking will require substantial resources, 

and encourages the Agency to develop and strengthen 

partnerships with appropriate federal and state agencies 

and private-sector organizations having parallel interests in 

advancing the necessary underlying science of Nr creation, 

transport and transformation, impacts, and management.

4.2. Consideration of Nr Impacts in Risk 

Reduction Strategies

Historical measurement and impact categories

The types of impacts of Nr in the environment are 

dependent on three general factors: the sources of Nr, the 

types of media impacted, and Nr chemical forms. The 

magnitude of effects depends on loading and the nature 

of the system impacted. As illustrated in Figure 1, the 

impacts of a given source of Nr can be multiple as N is 

transformed in the environment and transported among 

ecosystem components. The nitrogen cascade provides a 

comprehensive framework for understanding the role of 

Nr in the earth’s ecosystems and establishes a framework 

for developing and implementing management methods 

through which beneficial effects can be enhanced while 

minimizing detrimental impacts.

A management paradigm in which various approaches 

are used to limit environmental impacts to “acceptable” 

levels of risk is a useful concept for understanding the 

environmental impacts that Nr can have. For this purpose, 

impacts are divided into several general categories 

within which various contaminants have a direct 

correlation with damage. Once the nature and type of 

impacts are recognized, the risks should be characterized 

quantitatively, if possible. This information would then be 

used, along with other considerations such as economic, 

social and legal factors, to reach decisions regarding risk 

reduction strategies and the need for and practicability 

of implementing various risk reduction activities. The 

regulation of Nr in the environment by EPA follows an 

impact-by-impact approach which, with few exceptions, 

examines specific N forms in either aquatic, atmospheric, 

or terrestrial systems. As previously discussed, the 

principal regulatory authority pertaining to nitrogen is 

derived from the CWA and the CAA, although other 

legislation such as the Energy Independence and Security 

Act (EISA), and the Endangered Species Act (ESA) 

contain provisions that could result in regulatory actions 

that affect nitrogen management.

Historically, EPA environmental protection programs 

have addressed impacts of Nr such as climate change, 

eutrophication, ecotoxicity, human health (cancer 

and non-cancer), acidification, smog formation, and 

stratospheric ozone depletion, among others (Bare et al., 

2003). Within these categories it is sometimes possible 

to express end points in terms of collective metrics, such 

as is done with greenhouse gases in the form of carbon 

dioxide equivalents, or acidification as H+ equivalents. 

This approach has the considerable advantage of defining 

a straightforward framework within which environmental 

standards can be derived that are protective of human 

health and the environment – EPA’s principal mission. 

This approach also encourages evaluation of damage 
44

from multiple sources, as long as the characterization 

metric used is genuinely representative of the impact of 

a given contaminant. Thus, for example, the total impact 

of acidic gases such as SO2 and NOx on the acidification 

of watersheds can be expressed as a common metric. 

However, metrics for human health are generally not as 

simple to characterize nor are the appropriate end points; 

thus, the mechanism of toxicity, number of individuals 

affected, value of lost workdays, medical treatment costs, 

and value of human lives lost may all be used. 

Ecosystem functions and services

A complementary approach to classical impact 

characterizations is the use of ecosystem “service” and 

“function” categories, in which the impairment of a specific 

service provided by one or more ecosystems or impairment 

of an ecological function by causative contaminant 

emissions is assessed (Costanza, 1997; Millennium 

Ecosystem Assessment, 2003). Such an approach is 

inherently attractive because of its basis in scientific reality, 

i.e., the health of humans is inextricably linked to the health 

of the environment. Less clear, in some cases, are ways in 

which to measure and monitor such impacts and account 

for the effects of a complex array of factors and stressors 

that contribute to or damage ecosystem service, function, 

and health. Table 11 provides examples of ecosystem 

services and corresponding functions. 

The use of ecosystem services in a regulatory context 

would be a different approach for the EPA, one with 

considerable potential, but one for which experience is 

currently lacking. In comparison to the available data on 

reactive nitrogen usage, little is known about the response 

of ecosystems and ecosystem services to reactive nitrogen 

loads. This is discussed more fully in Sections 4.5, 4.6 and 

Appendix D on critical loads. In this context the Committee 

Table 11: Ecosystem service and corresponding function categories

Ecosystem Service Ecosystem Function

Gas regulation Regulation of atmospheric chemical composition

Climate regulation

Regulation of global temperature, precipitation, and other 

biologically mediated climatic processes at global, regional, 

and local levels

Disturbance regulation Capacitance, damping, and integrity of ecosystem response 

to environmental fluctuations

Water regulation Regulation of hydrologic flows

Water supply Storage and retention of water

Erosion control and sediment retention Retention of soil within an ecosystem

Soil formation Soil formation processes

Nutrient cycling Storage, internal cycling, processing, and acquisition of 

nutrients

Waste treatment Recovery of mobile nutrients, and removal or breakdown of 

toxic compounds

Pollination Movement of floral gametes

Biological control Trophic dynamic regulation of populations

Refugia Habitat for resident and transient populations

Food production That portion of gross primary production extractable as food

Raw materials That portion of gross primary production extractable as raw 

materials

Genetic resources Sources of unique biological materials and products

Recreation Providing opportunities for recreational activities

Cultural Providing opportunities for noncommercial uses

Source: Costanza et al., 1997 (Table 1, p. 254). Reprinted with permission; copyright 1997, Nature Publishing Group. 
45

supports plans by the EPA to incorporate research on 

the services concept, focusing on Nr as the suite of 

contaminants of interest, into its future ecological research 

plan (U.S. EPA, 2009a). EPA’s Ecological Research Plan 

was reviewed by the Science Advisory Board (U.S. EPA 

SAB, 2008). More recently, the Science Advisory Board 

completed a self-initiated study on “Valuing the Protection 

of Ecological Systems and Services” (U.S. EPA SAB, 

2009). This report explores the concept of ecosystem 

services as a basis for regulatory action and presents a 

roadmap for implementing this approach.

Economic measures and impacts

It is also possible to translate the effects of Nr into 

economic terms. Two economic measures that are often 

used are the dollar costs of damages and the cost of 

remediation or substitution. Another important economic 

metric is the cost/ton of remediation for each form of Nr. 

Damage costs do not always scale as tons of Nr released 

into the environment. If damage costs rather than tons of 

nitrogen were utilized as a metric, the full implications of 

the cascade and the setting of priorities for intervention 

might differ.

It is important to note that the choice of metric used in 

assessing impacts may play an influential role in what and 

how one manages. Air and water protection laws state that 

the goal is “to protect human health and the environment.” 

Yet, there is no generally agreed-upon common metric for 

measuring the full range of effects (which are complex 

and often unknown) or for setting priorities in the 

establishment or implementation of policies. 

As noted above, there are multiple metrics for 

measuring Nr or any other agent in the environment. The 

most common metric utilizes quantitative measures of 

the total amount of Nr (and any of its specific chemical 

forms) in different environmental reservoirs and the 

mass flux between them. But while providing common 

units, typically mass or concentration, these measures 

do not distinguish the relative societal costs of health 

or environmental consequences of reactive nitrogen of 

different forms or places in the cascade. While not all 

damages can be turned into economic costs, and the costs 

of some damages have not been quantified, enough of the 

major damages can be quantified economically to provide 

a useful complementary metric for decision-making. (See 

the Chesapeake Bay example in Box 2.)

The advantage of monetizing damages is that it reflects 

an integrated value that human society places on lost 

ecosystem goods and services in common currency and 

illustrates the cascading costs of damages as Nr changes 

form and moves between different parts of the ecosystem. 

In addition, human health implications can also be 

included as the cost of health care treatment, lost work 

days and other aspects of morbidity and mortality (e.g., 

economic value of lives lost). A third metric is to look at 

morbidity and/or mortality separately and not monetize 

them with a cost value. Of course a concern, particularly 

with respect to the economic metric, is that there are a 

number of ecosystem services that arguably cannot be 

easily monetized, for example the loss of biodiversity and 

those ecosystem functions that are affected by climate 

change or other stressors. Ecosystem services considered 

to be regulating and supporting are particularly difficult 

to fit into an economic metric. It is thus essential that 

a variety of complementary metrics be used to assess 

the impact of anthropogenic Nr on the environment and 

human well being. 

There is value in each of the ways that N metrics 

are expressed. Traditional categories provide a readily 

adaptable framework for regulation, while ecosystem 

service and function-based categories provide a richer 

context for stating the complex connections among 

Nr inputs and transformations and their impacts on 

ecosystem health and human well-being. Dollar-based 

metrics provide a means of identifying those effects that 

have the greatest impacts and costs to society. 

Finding 14: The Committee finds that reliance on only 

one approach for categorizing the measurement of Nr is 

unlikely to result in the desired outcome of translating 

N-induced degradation into the level of understanding 

needed to develop support for implementing effective Nr 

management strategies. 

Recommendation 14: It is, therefore, recommended 

that the EPA consider the impact of different metrics 

and examine the full range of traditional and ecosystem 

response categories, including economic and ecosystem 

services, as a basis for expressing Nr impacts in the 

environment, and for building better understanding and 

support for integrated management efforts.

4.3. Water Quality Regulation and 

Management

Aquatic thresholds

In aquatic ecosystems, thresholds at which excess Nr 

becomes a problem can be expressed as a management 

goal such as a total maximum daily load (TMDL) or as a 

critical load (CL). Under the authority of the CWA, EPA 

has developed guidance for establishing numeric nutrient 

criteria on an eco-regional basis for lakes and reservoirs, 

streams and rivers, estuaries and coastal waters, and 

wetlands. EPA has proposed specific numbers for lakes 

and reservoirs and rivers and streams and protocols for 

developing criteria for estuaries and wetlands. Each state is 

advised to go through an assessment to determine the best 

methodology for implementing numeric criteria (U.S EPA, 

2000c, 2000c, 2001b, 2007e). These criteria will identify 

impaired waterbodies for which TMDLs may be required.

The second type of threshold available for aquatic 

ecosystems is the critical load (CL). Unlike the TMDL, 

the CL (in the U.S.) has no regulatory framework but 

rather sets the threshold of Nr loading at which negative 
46

Box 2: Economic Impact and Metrics for Chesapeake Bay and Its Watershed

Recently, the N cycle and the implications of the reactive nitrogen cascade were translated into economic terms 

for the case of Chesapeake Bay (Moomaw and Birch, 2005). This approach has recently been updated with more 

recent data, and the economic and health impacts of different forms of Nr in multiple ecosystems and media have 

been estimated using better modeling methods (Birch et al., 2011). As an illustration, each of these metrics is shown 

as a percentage of Nr fluxes in the Chesapeake Bay water and air shed in Figure 15. Abatement costs are summarized 

in Table 12. Atmospheric emissions account for 37% 

of Nr entering the watershed, but they account for 

75% of the dollar damages and 76% of the mortality 

(U.S. EPA, 2005c). Mitigation costs per tonne of 

atmospherically released Nr are the lowest among the 

three sources. Additions of Nr to terrestrial ecosystems 

add 60% to the system, but contribute only 24% of the 

damage costs and 24% of the mortality, and have the 

highest mitigation costs. Freshwater releases, the second 

most expensive to mitigate, account for the smallest 

portion of Nr contributions to the system by any of the 

metrics considered: only 4% of the Nr, 2% of the cost 

damages, and none of the mortality losses (Birch et al., 

2011). Costs of Nr damage and health metrics provide 

additional economic measures of the cost effectiveness 

of actions to reduce a metric ton of Nr.

The metrics of damage cost and mortality (morbidity 

shows a similar pattern to mortality, but is only one-tenth 

the damage cost) indicate that controlling emissions 

of NOx from combustion and industrial processes 

produces greater gains in protecting human health and 

the environment than does reducing other Nr releases, 

though the two sources are comparable in terms of 

reactive forms of Nr released to the watershed. This 

difference occurs because emissions to the air cascade 

through more parts of the watershed ecosystem than 

releases directly to the Bay. If human health effects are 

monetized, then the economic gains are even greater 

from reducing atmospheric emissions (see Figure 16). 

Figure 16 is a scatter plot of all quantifiable damage 

costs (including health impacts) relative to metric tons 

of Nr showing the significant difference in emphasis 

of the two metrics. Note that direct additions to the 

environment from agriculture are about 370,000 tonnes 

Nr/year, and cause $1.7 billion worth of damage. 

Emissions of NOx from mobile sources represent only 

180,000 metric tons Nr/year but cause nearly $4.4 

billion in damages each year, of which $108 million 

is attributable to nitrate loading of the Chesapeake 

Bay, $3.9 billion to human morbidity and mortality, 

and the remainder to other forms of damage, such as crop and commercial forest damage. Hence the releases of Nr 

into the airshed from mobile sources, which are only half the amount of agricultural releases to the watershed, cause 

more than 2.5 times the economic damage of environmental additions from agriculture. This integrated inclusion of 

atmospheric, terrestrial, and aquatic additions of Nr is not reflected in today’s regulations

Marginal abatement costs per metric ton Nr by source into each of the three media are provided in Table 12, and 

demonstrate that the least costly abatement cost per metric ton of Nr also comes from atmospheric emission controls. 

While most legislation constrains how cost for remediation can be considered, it is useful to know where the lowest 

cost options lie in setting priorities.

Figure 15: Relative importance of all reactive 

nitrogen sources released into atmospheric, 

terrestrial, and freshwater media within the 

Chesapeake Bay Watershed utilizing four 

different metrics 

Source: Birch et al., 2011 (Figure 3, p. 173). Reprinted 

with permission; copyright 2011, American Chemical 

Society. 

Figure 16: Quantified damage costs (including 

health impacts) relative to metric tons of reactive 

nitrogen in Chesapeake Bay Watershed

Source: Birch et al., 2011 (Figure 4, p. 173). Reprinted 

with permission; copyright 2011, American Chemical 

Society.
47

impacts have been documented. Based extensively on 

European work, CLs for aquatic ecosystems are Nr inputs 

on the order of 2-15 kg N/ha/yr (Bobbink et al., 2010). 

There are numerous locations within the U.S. where 

deposition to surface waters falls within this range.

Water quality standards

Section 303 of the CWA requires states to adopt 

water quality standards and criteria that meet the stateidentified designated uses (e.g., uses related to “fishable” 

and “swimmable”) for each waterbody. Specifically, “a 

water quality standard defines the water quality goals of 

a water body, or portion thereof, by designating the use 

or uses to be made of the water and by setting criteria 

necessary to protect the uses” (40 CFR § 131.2). Further, 

“such standards serve the dual purposes of establishing 

the water quality goals for a specific water body and serve 

Table 12: Marginal abatement cost per tonne of Nr by source

Location in the N cascade 

where emitted Source/pollutant Abatement cost per tonne of Nr

Air

Electric utilities/NOx 14 $4,800

Industrial/NOx 15 $22,000

Mobile sources/ NOx 16 $14,000

Non-agricultural/NH3 No estimate

Land

Agriculture/nitrate 17 $10,000

Urban and mixed open land 

uses/nitrate 17 $96,000

Fresh water Point sources/nitrates 17 $18,000

These multiple metrics provide several ways of looking at the nitrogen cascade and its impact on human health 

and the environment. However, there are many impacts that remain unaccounted for in any of these metrics. Some 

impacts might be quantified, but the necessary data have yet to be collected. Economic losses due to damage to 

commercial fisheries in the Bay are an example that is likely to be significant but has not yet been quantified. 

Similarly, economic losses due to climate change and ozone depletion from N2O emissions have not been fully 

evaluated. Impacts such as loss of biodiversity cannot be readily quantified at all, so it is desirable to consider a set of 

qualitative and non-quantified metrics in addition to the quantitative ones.

Other parts of the country such as the Mississippi valley or the Central Valley of California are expected to show 

very different patterns of cost damages, with terrestrial and freshwater emissions causing proportionally higher 

damage costs, and emissions to the atmosphere causing a lower percentage of damages. But those very differences 

would assist EPA and the generators of those emissions in setting priorities for mitigation.

It is important to recognize that Nr is not the only stressor that can affect both human and environmental health. 

Researchers are challenged to comprehensively understand cause-and-effect relationships in a complex environment 

and to balance management actions and costs to ensure that risk-minimizing management strategies are effectively 

implemented.

As these multiple metrics indicate, decisions about which fluxes of Nr to mitigate depend upon which metric is 

utilized. The cascading economic costs of damage highlight the importance of regulating air emissions because of 

their impacts on human heath as well as their large contribution to the degradation of Chesapeake Bay water quality. 

Hence, if one is interested in reducing water impacts of Nr, the total reduction of damage may rely nearly as much 

on stricter enforcement of the Clean Air Act as the Clean Water Act. This challenges our traditional approach to 

regulation, but that is a consequence of comprehensively examining Nr guided by the nitrogen cascade.

14 See U.S. EPA, 2005c

15 See U.S. EPA, 1998

16 See Krupnick et al., 1998

17 See Chesapeake Bay Program, 2003a,b
48

as the regulatory basis for the establishment of water 

quality-based treatment controls and strategies beyond the 

technology-based levels of treatment required by sections 

301(b) and 306 of the Act” (40 CFR § 131.2).

The EPA sets minimum requirements for approvable 

standards and criteria including: use designations; 

water quality criteria sufficient to protect the designated 

uses; and an antidigradation policy (40 CFR § 131.6). 

Traditionally, Nr and other land, air, and water pollutants 

are measured in terms of quantity (mass) released per unit 

time (e.g., kg/day) or as a concentration (e.g., milligrams 

per liter, ml/L). Therefore, regulations often specify mass 

loading limits or maximum concentrations in permits.

In the mid-to-late 1990s, EPA began to emphasize the 

development of numeric nutrient criteria for both P and 

N through the state standards-setting process because, 

according to the 1996 Water Quality Report to Congress 

(U.S.EPA, 1997), 40% of the rivers, 51% of the lakes and 

ponds, and 57% of the estuaries assessed for the report 

were exhibiting a nutrient-related impairment. Few states 

had adopted numeric nutrient criteria for all affected 

waterbodies, especially for N, often relying on narrative 

criteria or secondary effects such as chlorophyll-a

concentration, dissolved O2, or water clarity. EPA’s 

strategy, driven by President Clinton’s Clean Water 

Action Plan (U.S. EPA and USDA, 1998) mandated 

numeric nutrient criteria to begin to address the problem 

(U.S. EPA, 1999). To move the objectives of the Clean 

Water Action Plan forward, EPA published national 

nutrient criteria guidance for lakes and reservoirs (U.S. 

EPA, 2000d), rivers and streams (U.S. EPA, 2000b), 

estuaries and coastal waters (U.S. EPA, 2001b), and 

wetlands (U.S. EPA, 2007c), based upon ecoregional 

guidance for lakes and reservoirs and rivers and streams. 

To date, relatively few states have adopted new numeric 

criteria into their water quality standards. While some 

successes are evident in promulgating P criteria for 

freshwater systems, which has a richer history of numeric 

criteria incorporation into state water quality standards, 

development of numeric nitrogen criteria has been elusive 

for a variety of reasons. 

Nr management in multiple media and across 

jurisdictions can be complicated because the CWA has 

little authority over atmospheric sources, and individual 

states explicitly lack authority to control upstream 

sources. For example, extensive monitoring and analysis 

of the sources of reactive nitrogen in the Raccoon 

River of western Iowa have shown that point sources 

from municipal treatment plants and residential septic 

tanks account for less than 8% of the total nitrogen 

load to the system, with agricultural runoff being the 

overwhelming source (Jha et al., 2010). This disparity is 

similar statewide (Libra et al., 2004). As a result, nutrient 

management strategies that are focused on the control 

of point sources can often result in inefficient allocation 

of resources if non-point sources are not also addressed. 

In addition it is often the case for estuaries such as the 

Gulf of Mexico or Chesapeake Bay, that management 

goals that meet water quality standards cannot be attained 

without interstate compacts or a strong federal role. This 

may be resisted by upstream states that may have to bear 

the cost but do not necessarily reap the benefits of the 

water quality improvement. Such a dilemma underscores 

the need for an integrated approach to Nr management. 

The Committee notes that a State-EPA Nutrient 

Innovations Task Group has considered some options for 

improving control of nutrient pollution sources (StateEPA Nutrient Innovations Task Group, 2009).

Populated (urban/suburban/developed) land areas 

provide significant loads of Nr to the environment, both 

by generation (e.g., deposition of NOx emissions) and 

by transfer (e.g., domestic sewage from food imported 

into the watershed). Categorical sources include sewage 

treatment plants (STPs), industries, subsurface (septic) 

systems, atmospheric deposition, domestic animal and 

wildlife waste, and fertilizers used on lawns, gardens 

and landscapes. Infrastructure (e.g., storm sewers) and 

landscape conditions (e.g., increased impervious cover) 

more efficiently move Nr associated with surface runoff 

to receiving waters and may also inject or infiltrate Nr into 

ground water. Landscape changes, primarily increases in 

impervious cover, soil disturbance and compaction, and 

wetland/hydric soil losses, have also reduced the capacity 

for natural systems to treat Nr inputs by recycling or 

denitrification. Other disruptions in chemical condition 

(e.g., acidification), biology (e.g., vegetative cover), and 

physical character (e.g., temperature increase) alter the 

nitrogen cascade, which may have both negative and 

positive consequences for Nr amelioration on the populated 

landscape and in air and water. Populated lands are 

estimated to export as much as 10 times the total nitrogen 

that was exported under pre-development conditions. 

Finding 15: Intervention to control Nr under most 

water management programs generally occurs in three 

ways: 

n Prevention or source controls.

n Physical, chemical, or biological “dead ending” or 

storage within landscape compartments where it is 

rendered less harmful (e.g., long-term storage in soils 

or vegetation; denitrification, primarily in wetlands; 

reuse).

n Treatment using engineered systems such as wastewater 

treatment plants or BMPs for stormwater and nonpoint 

source runoff. 

While most management programs focus on the 

third (treatment) approach, there are opportunities for 

combining the three that can be more effective and cost 

less. Furthermore, it is important to recognize that in 

some cases total reduction of water impacts of Nr may 

rely nearly as much on stricter enforcement of the Clean 

Air Act as the Clean Water Act.
49

Recommendation 15: To better address Nr runoff 

and discharges from the peopled landscape the Committee 

recommends that EPA:

15a.

• Evaluate the suite of regulatory and non-regulatory 

tools used to manage Nr in populated areas from 

nonpoint sources, stormwater and domestic sewage, 

and industrial wastewater treatment facilities, 

including goal-setting through water quality standards 

and criteria. 

• Determine the most effective regulatory and voluntary 

mechanisms to apply to each source type (recognizing 

that in some cases total reduction of the impacts of Nr 

may rely nearly as much on stricter enforcement of 

the Clean Air Act as the Clean Water Act) with special 

attention to the need to regulate nonpoint source and 

related land use practices.

15b.

• Review current regulatory practices for point sources, 

including both wastewater treatment plants and 

stormwater, to determine adequacy and capacity 

towards meeting national Nr management goals. 

• Consider technology limitations, multiple pollutant 

benefits, and funding mechanisms as well as potential 

impacts on climate change from energy use and 

greenhouse gas emissions, including nitrous oxide.

15c.

• Set Nr management goals on a regional/local basis, 

as appropriate, to ensure most effective use of limited 

management dollars. 

• Fully consider “green” management practices such as 

low- impact development and conservation measures 

that preserve or re-establish Nr-removing features to the 

landscape as part of an integrated management strategy, 

along with traditional engineered best management 

practices.

15d.

• Research best management practices that are effective in 

controlling Nr, especially for nonpoint and stormwater 

sources, including land and landscape feature preservation 

and set Nr management targets that realistically reflect 

these management and preservation capacities. 

• Construct a decision framework to assess and determine 

implementation actions consistent with management goals.

15e.

• In cooperation with the Departments of Agriculture and 

Army, the Fish and Wildlife Service and the Federal 

Emergency Management Agency, the EPA should 

develop programs to encourage wetland restoration 

and creation with strategic placement of these wetlands 

where reactive nitrogen is highest in ditches, streams, 

and rivers. The Agency should also address the means 

of financing, governance, monitoring, and verification. 

Such programs might be modeled on the Conservation 

Reserve Program or extant water quality and 

environmental trading programs, but need not be limited 

to current practices (as discussed in section 5.3.4).

4.4. Water Quality Monitoring 

and Assessment

Under Section 106 of the CWA, the EPA provides 

funds to assist state and interstate agencies and tribes 

to conduct monitoring of the nation’s waters to ensure 

adopted water quality criteria and designated uses are 

met. Further, primarily under Section 305(b) of the CWA, 

those entities are required to report, on a biennial basis, on 

the health and status of their jurisdictional waters. These 

assessments are presented by the states to the EPA to 

categorize attainment of designated uses. EPA published 

these reports up until 1998 (U.S. EPA, 2000a), after 

which it transitioned into a Water Quality Report in 2000 

(U.S. EPA, 2002) and a National Assessment Database 

in 2002 (U.S. EPA, 2010c). States also prepare a list of 

“impaired” waters under Section 303(d) of the CWA and 

EPA develops a synthesis of the CWA Section 305(b) and 

303(d) reporting under a Consolidated Assessment and 

Listing Methodology (CALM) approach.

As discussed above, the EPA compiles the approved 

state 303(d) lists into a national listing (U.S. EPA, 

2010e). The list provides information by state as well 

as by impairment cause, and identifies the TMDLs 

completed to date. The most current data available 

on the EPA Web site includes reporting from most 

entities through 2008. The report identifies 6,816 

impairments related to “nutrients” (almost 9% of all 

identified impairments), although other impairments 

may ultimately have a nutrient enrichment cause. For 

example, organic enrichment/oxygen depletion (6,410), 

turbidity (3,046), noxious aquatic plants (981), algal 

growth (539), and ammonia (general toxicity 356), can 

all have a common cause such as N or P enrichment. It 

should also be clear that impairments may have multiple 

causes so, for example, waters identified as impaired by 

O2 depletion may also be impaired by nutrients.

There are other initiatives promoted by EPA to 

monitor and assess the nation’s waters, generally 

implemented in collaboration with, or by, the state and 

interstate agencies and tribes having jurisdiction over the 

waters. These include the Wadeable Stream Assessment 

(U.S. EPA, 2006c), the National Coastal Assessment 

and its National Coastal Condition Reports (U.S. EPA, 

2001a, 2004a, 2006b), the Survey of the Nation’s Lakes 

and Survey of the Nation’s Rivers and Streams, and 

more recently, probabilistic monitoring efforts in lakes, 

streams, and estuaries (U.S. EPA, 2010d). Many of 

these are aimed at including a biological assessment 

component that is often lacking in water pollutant and 

chemistry efforts described above.

The USGS collects data on surface and underground 

waters and disseminates these data to the public, state and 

local governments, public and private utilities, and other 
50

federal agencies involved with managing water resources. 

The Committee encourages EPA to work closely with 

USGS on monitoring and assessment activities.

The National Oceanic and Atmospheric Administration 

has periodically produced estuarine assessments under 

the National Estuarine Eutrophication Assessment 

(NEEA) program. The most recent report was released 

in 2007 (Bricker et al., 2007). The report has a focus on 

nutrient enrichment and its manifestations in the estuarine 

environment and relies on participation and interviews of 

local experts to provide data for the assessment. Among 

the key findings were: 

n Eutrophication is a widespread problem, with the 

majority of assessed estuaries showing signs of 

eutrophication – 65% of the assessed systems, 

representing 78% of assessed estuarine area, had 

moderate to high overall eutrophic conditions. 

n The most common symptoms of eutrophication were 

high spatial coverage and frequency of elevated 

chlorophyll-a (phytoplankton) – 50% of the assessed 

estuaries, representing 72% of assessed area, had 

excessive chlorophyll-a ratings.

4.5. Clean Air Act and Air Quality 

Regulation and Management 

The modern history of American air pollution control 

legislation begins with the 1963 Clean Air Act (CAA) 

which, along with its amendments, requires the EPA 

to establish and revise National Ambient Air Quality 

Standards (NAAQS) and to prepare state of the science 

reviews such as the Criteria Documents and more recently 

the Integrated Science Assessments (ISA) (U.S. EPA, 

2005a, 2006a, 2007a). There are six criteria pollutants: 

carbon monoxide, lead, NO2, ozone, SO2, and PM. 

These have been determined to endanger public health 

or welfare. The CAA as currently written requires a 

review of the scientific criteria for these standards at fiveyear intervals. Although NO2 is the only Nr compound 

specified as a criteria pollutant, NHx and NOy play a 

major role in formation of the secondary pollutants ozone 

and particulate matter. 

The CAA has been amended several times since its 

inception. In 1970, the CAA was amended “to provide 

for a more effective program to improve the quality of 

the nation’s air.” The CAA was amended again in 1977, 

primarily to mandate reductions of emissions from 

automobiles. Despite evidence that NOx is the central 

pollutant in photochemical smog formation (Chameides 

and Walker, 1973; Crutzen, 1973, 1974; Fishman and 

Crutzen, 1978; Fishman, et al., 1979), federal regulations 

did not require automobiles to control NOx emissions to 

below 1 g/mi (0.14 g N per km) until 1981. Few locales 

violate the standards for NO2,

18 but the secondary effects 

of several of these gases also pose health and welfare 

concerns. If a city had an annual average NO2 level 

Table 13: Federal primary ambient air quality standards that involve Nr, effective February 2010.

Pollutant Federal Primary Standard (NAAQS)

Ozone (O3)

 1-hr average

 8-hr average

0.12 ppmv

0.08 ppmv

Nitrogen Dioxide (NO2)

 1-hr average

 Annual average

100 ppb

0.053 ppmv (100 μg/m3)

Particulate Matter, coarse (PM10)

 Diameter ≤ 10 μm, 24-hr average

 Annual average

Particulate Matter, fine (PM2.5)

 Diameter ≤ 2.5 μm, 24-hr average

 Annual average

150 μg/m3

50 μg/m35 μg/m3

15 μg/m3

Note: Secondary standards are currently identical to the primary standards. Source: www.epa.gov/air/criteria.html

18 In 2010, EPA promulgated a new 1-hour standard of 100 ppb for NO2 [Primary National Ambient Air Quality Standards for Nitrogen Dioxide; Final 

Rule, Federal Register 75 (26): 6474-6537]. Monitoring for compliance with this new standard is required, but it will not be known for several years 

which if any locales violate this standard.
51

anywhere near the NAAQS for NO2, it would risk severe 

photochemical smog – the summertime efficiency for 

ozone production ranges from 4 to 10 ppb O3 per ppb NOx. 

As previously discussed, the focus on compliance 

monitoring for NO2 ignores the other, equally important 

members of the NOy family such as HNO3 that deposits 

quickly onto the earth’s surface. It is clear that a causal 

relationship exists between current levels of N and S 

deposition and numerous biologically adverse effects 

on ecosystems across the U.S. (U.S. EPA, 2008d). 

Conversion of the existing network of NOx monitors to 

NOy monitors with a detection limit of 0.1 ppb would 

still demonstrate compliance with the NO2 standard but 

greatly increase the utility of the measurements for model 

evaluation as well as for understanding nitrate deposition 

and formation of photochemical smog, and haze.

Air pollution, especially ozone and PM, continued to 

be a problem in many American cities and the CAA was 

again amended in 1990. The Nr-relevant aspects were 

aimed at controlling urban smog and acid deposition. 

States were required to develop emissions inventories for 

reactive organic compounds, carbon monoxide, and NOx, 

but not NH3 or N2O. Over the U.S., sulfate and nitrate are 

responsible for about two-thirds and one-third, respectively, 

of the direct deposition of acids. The CAA Amendment of 

1990 required emissions decreases of 10 million tons of SO2

and 2 million tons of NOx relative to 1980 levels. Ammonia 

and ammonium, although they contribute to acidity after 

entering terrestrial ecosystems (Galloway et al., 2003; NRC, 

2003) and are expected to play an increasing role (Pinder et 

al., 2008), were not regulated by this legislation. 

The 1997 revision of the CAA and related regulations 

changed the standards for ozone and PM (see Table 13). A 

sizable fraction of the mass of PM less than 2.5 microns, 

PM2.5, is condensed Nr. As stated above, these particles 

have adverse health consequences. PM is also controlled 

by the Regional Haze Regulations (40 CFR 51). These 

regulations require that by the year 2064, states must 

restore Class I areas defined in the regulations to their 

natural levels of atmospheric clarity. 

Ozone and PM, the two most recalcitrant of the criteria 

pollutants, cover large spatial scales. All of the ozone and 

much of PM are secondary pollutants in that they are not 

released at the tailpipe but form in the atmosphere. Ample 

evidence shows that much or most of the PM in American 

cities is secondary (e.g., Donahue et al., 2009). Violations 

are declared on urban scales, responsibility for their control 

was assigned to states, but the physics and chemistry of 

smog and haze are regional. In the eastern U.S., ozone 

episodes often cover several states and involve pollutants 

emitted in upwind states that do not themselves experience 

violations (Husar et al., 1977; Logan 1989; Moy et al., 

1994; Ryan et al. 1998). The 1990 amendments to the Clean 

Air Act authorized, in part as a response to this scaling 

problem, the Ozone Transport Assessment Group (OTAG) 

and the Ozone Transport Commission (OTC). These have 

jurisdiction extending from Washington, D.C. to Maine. 

Progress has been made on regional control of emissions; 

the NOx State Implementation Plan (SIP) call, implemented 

in 2003 and 2004, has led to measurable improvements in 

ambient ozone and nitrate levels (Gego et al., 2007; Sickles 

and Shadwick, 2007a). Experiences with ozone and PM 

provide a useful demonstration of why it is necessary to 

develop an integrated approach to management of Nr.

Atmospheric thresholds for Nr 

As shown in Table 13 the metric used for safe, upper 

limits in the atmospheric environment is concentration 

(in mass per unit volume of air or volume mixing ratios) 

averaged for a given time period, usually 1 hour, 8 hours, 

24 hours, or annually. The thresholds for excess Nr in the 

atmosphere remain an area of active research. The only 

Nr compound for which there is currently a NAAQS is 

NO2, which may not exceed 0.053 ppm (100 μg/m3) for 

the annual arithmetic mean and 100 ppb for the one-hour 

average. This standard, based on the direct health effects, 

is certainly inadequate because NO2 concentrations well 

below 0.053 ppm lead to concentrations of secondary 

pollutants well above acceptable levels (i.e., PM2.5 and O3). 

The NO2 concentration required to achieve the current 75 

ppb ozone standard has not been rigorously established, 

but it must be well below 0.053 ppm, because information 

provided by EPA indicates that areas currently in violation 

of the ozone standard typically have NO2 concentrations 

below 0.020 ppm (U.S.EPA, 2010a). The NO2 concentration 

required to achieve the current 15 μg/m3 PM2.5 standard 

is probably also below the 100 μg/m3 standard for NO2

because of the role of NO2 in secondary particulate 

formation. States in the eastern U.S. are considering 

substantial additional NOx emissions reductions in order to 

comply with the new 8-hour 75 ppb ozone standard. One 

scenario being tested (G. Aburn, Maryland Department 

of Environment, personal communication) involves the 

following reductions: (1) reducing NOx emissions for point 

sources by 65%, (2) reducing NOx emissions for on-road 

sources by 75 percent, (3) reducing NOx emissions for 

nonroad sources by 35%, and (4) reducing VOC emissions 

by 30% for all source groups.

As further discussed in Section 6.2, it is the opinion of 

the Committee that a decrease in NOx emissions of 2 Tg N/

yr relative to the 2002 baseline level can be achieved in the 

near term. Emissions decreases implemented since 2002 

have already substantially improved ozone concentrations 

(Gégo et al., 2007). The absolute amount of decrease and 

the positive impact it would have on human health is region 

dependent, but further decreases will result in further 

beneficial decreases in PM2.5 and O3 concentrations. 

The threshold for total Nr in the atmosphere is yet to be 

fixed, but depends on its rate of deposition to the surface 

and the sensitivity of the receptor(s). The immediate 

need for determining thresholds for atmospheric Nr is 

monitoring of NOy and NHx.
52

4.6. Thresholds for Excess Nr Effects on 

Terrestrial Ecosystems

In parallel with the original concept of critical loads 

developed by Nilsson and Grennfelt in 1988 and now 

widely used for air quality management in Europe 

(Appendix D), thresholds in general and critical loads 

specifically for Nr effects on terrestrial ecosystems in the 

U.S. should be understood to be “quantitative estimates 

of exposure to air concentrations of Nr compounds below 

which harmful effects on specified sensitive elements 

within ecosystem of concern do not occur according 

to present knowledge” (Nilsson and Grennfelt, 1988; 

Heittelingh et al., 2001). 

In developing these quantitative estimates of thresholds 

and/or critical loads for terrestrial ecosystems in the U.S. 

(e.g., Fenn et al., 2003), it is imperative to understand 

the extraordinarily wide diversity and Nr-sensitivity 

of various components of terrestrial ecosystems in 

different parts of the U.S., as well as the huge differences 

in purposes and intensity of management and public 

perceptions of the value of these ecosystem components 

to various sectors of American society. Thus, the critical 

loads appropriate for maintaining species diversity in a 

natural grassland in northern Minnesota or a wilderness 

area in the Mediterranean climate of southern California 

are likely to be very different from those associated 

with direct effects on similar systems in other regions 

of the U.S. – or even for beneficial and/or adverse 

effects on other components of the same terrestrial 

ecosystem. For example, the threshold or critical load for 

adverse effects of excess Nr on understory vegetation, 

beneficial mycorrhizae, or lichen communities in a 

forest ecosystem is likely to be very different from the 

threshold for adverse effects on the dominant forest trees 

in that same ecosystem. Thus, public perceptions of 

“specified sensitive elements within the ecosystem” may 

be important in determining what specific thresholds or 

critical loads should be considered in order to minimize or 

avoid specific adverse effects of concern.

At present, the sum total of directly measured wet 

plus dry-deposited chemically oxidized (NOy) and 

chemically reduced (NHx) inorganic Nr loads in various 

states within the contiguous U.S. are on the order of 3 

to 15 kg N/ha/year (NADP, 2010; CASTNET, 2010). As 

shown in Appendix A, a three-year run of the Community 

Multiscale Air Quality (CMAQ) model also provided 

estimates of the average annual total Nr loads (including 

organic forms as well as inorganic NOy and NHx forms of 

Nr) in the contiguous U.S. These model estimates varied 

from minimal deposition values of about 3 kg N/ha/year to 

maximum estimated values of about 17 kg N/ha/year. This 

range agrees well with the range of the measurements. 

These directly measured and modeled estimates of total 

(wet plus dry) deposition of organic and inorganic forms 

of Nr indicate that there are several areas, especially in 

the eastern U.S., and a few areas of the western U.S., 

where current total Nr loads are already very close to, or 

will very likely soon exceed, the recommended threshold 

and critical load estimates provided by Bobbink et al. 

(2010) in their review of scientific evidence regarding 

the impacts of atmospheric nitrogen deposition on plant 

diversity in terrestrial ecosystems.

4.7. Comments on Nr Critical Loads

In recent years, the Acid Rain Action Plan developed 

by New England governors and eastern Canadian 

Premiers has led to evaluations of critical loads to surface 

waters and forests in that region. Those studies identified 

many waters and forest lands that met or exceeded 

critical load capacity for combined sulfur and nitrogen 

deposition both in the New England States and in the 

eastern Canadian provinces. The plan set target decreases 

of 20 to 30% for nitrogen oxide emissions by 2007 and a 

50% decrease in sulfur dioxide emissions by 2010. These 

targets are intended to decrease long-range transport of 

air pollutants, acid deposition, and nutrient enrichment of 

marine waters in this region.

In May 2006, a Multi-Agency Critical Loads Workshop 

was held, which led to the formation of a Critical 

Loads Ad-Hoc Committee (CLAD) within the National 

Atmospheric Deposition Program (NADP). A goal of 

the program is to “provide consistency in development 

and use of critical loads in the U.S.” One outcome is a 

project undertaken by the Northeast States for Coordinated 

Air Use Management (NESCAUM) to: “estimate critical 

loads of sulfur and nitrogen in atmospheric deposition 

for areas where sufficient knowledge, data, and methods 

exist” and “to demonstrate the use of critical loads as a tool 

for assessing environmental policies and programs and 

managing natural resources.”

A February 2007 Workshop sponsored by EPA on 

“The Assessment of Health Science for the Review of the 

National Ambient Air Quality Standards (NAAQS) for 

Nitrogen (NOx) and Sulfur Oxides (SOx)” expansively 

reviewed both ecosystem and human health effects toward 

revision of the NAAQS. Policy discussions at this workshop 

raised the questions of whether critical loads assessments 

were an effective means of improving ecosystem 

management, and whether the science was understood well 

enough to use critical loads as a management tool. The 

conclusion was that, although there was a substantial body 

of accumulated scientific evidence, there was only limited 

use of critical loads approaches for management of air 

quality in the U.S. The Multi-Agency Workshop on Critical 

Loads (mentioned above) was cited at EPA’s 2007 workshop 

as an agenda-setting effort to resolve some of the science 

and policy issues that could help advance critical loads 

approaches in the U.S. The Integrated Nitrogen Committee 

believes that the primary reason critical loads are not now 

used in the U.S. is that policy makers in this country have 

so far not been willing to adopt unfamiliar air and water 

quality management approaches or approaches that have not 
53

been evaluated directly in this country. Thus, the Committee 

recommends that EPA consider implementation of the 

critical loads concept for management of deleterious Nr 

effects in various parts of the U.S.

Finding 16: The Committee finds that there have been 

persistent increases in the amounts of Nr that have been 

emitted into and retained within various ecosystems, 

affecting their functioning. Unless this trend is reversed, 

it will become increasingly difficult for many of these 

ecosystems to provide the services upon which human 

well-being is dependent. The Committee believes that 

there is a need to regulate certain forms of Nr to address 

specific problems related to excess Nr, and we believe that 

the best approach for an overall management strategy is 

the concept of defining acceptable total Nr critical loads 

for a given environmental system.

Recommendation 16: The Committee recommends 

that the Agency work toward adopting the critical loads 

approach concept in determining thresholds for effects 

of excess Nr on terrestrial and aquatic ecosystems. 

In carrying out this recommendation the Committee 

recognizes that it will in many cases be necessary for 

the Agency to enter into new types of research, policy, 

and regulatory agreements with other federal, state, and 

tribal units based on cooperative, adaptive, and systemic 

approaches that derive from a common understanding of 

the nitrogen cascade. 

4.8. Tradeoffs of Nr Impacts in Risk 

Reduction Strategies

Because nitrogen is such an abundant and widespread 

element, and Nr is such a critical component of the earth’s 

biosphere, associated impacts are many and pervasive. 

In many cases, strategies to manage the impacts of Nr 

involve tradeoffs, i.e., mitigating one type of impact may 

exacerbate others. Given the interactions among oxidized 

and reduced N species, it is important to recognize the 

potential for unintended consequences to occur as a result 

of strategies that are aimed at limiting one form of Nr in air 

or water but lead to the increased production of other forms 

of Nr, or the formation and release of other contaminants 

of concern. For example, stringent control of point sources 

of Nr can be energy-intensive, requiring significant energy 

investments for chemicals, electricity, and other support, 

and this may in turn lead to the production of more reactive 

nitrogen and increased CO2 emissions. Furthermore, 

there may be environmental impacts of these treatment 

processes, particularly in the production of solid wastes 

that can be significant environmental hazards. This is 

the main reason that a life cycle approach is necessary 

in evaluating any remediation or treatment scheme. In 

addition, as discussed in Section 3.1.2, numerous lakes, 

reservoirs, rivers, and fjords worldwide exhibit N and P colimitation, either simultaneously or in seasonally-shifting 

patterns. Therefore, strategies are needed to reduce both 

P and N inputs. Not all control practices will be effective 

for dual nutrient reduction and this must be taken into 

consideration. Four categories of tradeoffs examined below 

are: ammonia release from concentrated feed lot operations 

(CAFOs), concerns about human nutrition, nitrification and 

denitrification, and nitrogen-carbon related impacts. 

Ammonia release from CAFOs

As a result of effluent guidelines for NH3 in aquatic 

systems, state and federal regulations and programs 

under the CWA were developed to address water 

quality protection from CAFOs. The resulting manure 

management systems utilized NH3 volatilization as a 

means to remove N and decrease the N in the manure 

when land applied. Only recently has the resulting 

increase in NH3 emission into the air been viewed as a 

potential problem with respect to air quality concerns and 

N deposition.

Finding 17: Current EPA policy (40 CFR Part 51, 

Clean Air Fine Particle Implementation Rule) discourages 

states from controlling ammonia emissions as part of 

their plan for reducing PM2.5 concentrations. In this 

rulemaking, EPA states that “ammonia reductions 

may be effective and appropriate for reducing PM2.5

concentrations in selected locations, but in other locations 

such reductions may lead to minimal reductions in PM2.5

concentrations and increased atmospheric acidity.” 

Ammonia is a substantial component of PM2.5 in most 

polluted areas of the United States at most times. While 

it is true that reducing NH3 emissions might increase the 

acidity of aerosols and precipitation, the net effect of NH3

on aquatic and terrestrial ecosystems is to increase acidity. 

After being deposited onto the earth’s surface, NH4

+ is 

under most circumstances quickly nitrified, increasing the 

acidity of soils and waters. The Committee is unaware of 

any evidence that NH3 reduces the toxicity of atmospheric 

aerosols or that high concentrations of NH3 occur 

naturally over any substantive area of the United States. 

It has not yet been established which components of PM 

have substantive impacts on human health, but the total 

concentration of PM2.5 correlates with morbidity and 

mortality, and NH3 contributes to PM2.5. The visibility 

degradation and other adverse effects associated with 

PM2.5 are related to aerosol surface area or mass where 

NH4

+ certainly plays a role. 

Recommendation 17: The Committee recommends that 

the EPA presumption that NH3 is not a PM2.5 precursor 

should be reversed and states should be encouraged to 

address NH3 as a harmful PM2.5 precursor.

Swapping N between environmental systems

Nitrous oxide is produced in “natural” and agricultural 

soils, and all aquatic systems, almost exclusively as 

a result of the microbial processes of nitrification and 

denitrification. As NH4

+ ion is the initial mineral N 

product formed during organic matter mineralization 

and most of the fertilizer used worldwide is NH4

+ based 

(e.g., urea, ammonium sulfate) (FAO, 2007), the suite 
54

of microbiological reactions that result in the release of 

gaseous N products need to be considered.

Nitrification is the oxidation of NH4

+ ion to 

NO3

-

 (Figure 17). Most commonly, nitrification is a 

chemolithotropic process consisting of the conversion of 

ammonium to nitrite, which is then converted to NO3

-

 by 

a second group of bacteria. The ammonium oxidizing 

bacteria (AOB) are obligate aerobes with some species 

that are tolerant of low-oxygen environments. The most 

common genera of autotrophic NH4

+ oxidizers are 

Nitrosospira and Nitrosomonas. AOB are found in most 

aerobic environments where ammonium is available 

through the mineralization of organic matter or where N 

compounds are added.

Biological denitrification is the dissimilatory reduction 

of NO3

-

 and nitrite to produce NO, N2O, and N2 by a 

taxonomically diverse group of bacteria. These bacteria 

synthesize a series of reductases that enable them to 

utilize successively more reduced N oxides as electron 

acceptors in the absence of oxygen. The general reductive 

sequence is shown in Figure 17. In addition to the 

free-living denitrifiers, symbiotically-living Rhizobia in 

root nodules of legumes are able to denitrify nitrate and 

produce nitrous oxide (Mosier and Parkin, 2007).

The abundant denitrifiers are heterotrophs, which 

require sources of electron-reducing equivalents contained 

in available organic matter. Factors that most strongly 

influence denitrification are oxygen, nitrate concentration, 

pH, temperature, and organic carbon. The reductive 

enzymes are repressed by oxygen but not by NH4

+. 

Nitrous oxide reductase appears to be more sensitive to 

oxygen than either NO3

-

 or nitrite reductase. Therefore N2

production predominates in more anoxic sites and N2O 

production may be greater in more aerobic conditions. 

However, the ratio of N2 to N2O emitted may also be 

affected by high NO3

-

 concentrations and associated higher 

levels of electrical conductivity and osmotic stress and soil 

pH (low pH favors N2O production). 

Given these interactions among oxidized and 

reduced N species (discussed above), it is important to 

recognize the potential for unintended consequences 

to occur as a result of strategies that may be aimed at 

limiting one form of Nr in air or water but lead to the 

increased production of other forms of Nr. One such 

instance is the potential offsetting of the benefits of 

NO3

-

 remediation at the expense of increasing input of 

N2O to the atmosphere. An example of such a situation 

involves NO3

-

 leached from agricultural fields, much of 

which could be removed from drainage water in natural 

or reconstructed wetlands. This process is ideal if the 

denitrification process goes to completion, i.e., only 

N2 is produced. If, however, the process is incomplete, 

and NO and N2O gases are emitted, then the end result 

may create a compensating risk that could be greater 

than that posed by the nitrate that is removed. This is 

because NO continues to be reactive in the atmosphere 

and is eventually redeposited in aquatic or terrestrial 

systems, and N2O is a GHG that has an atmospheric 

life time of approximately 100 years and a radiative 

forcing of approximately 300 times that of CO2 on a 

hundred-year time frame (IPCC, 2001). N2O is also a 

major source of NO in the stratosphere and depletes 

stratospheric ozone (Crutzen, 1981). If more of the NO3

-

denitrified is converted to N2O in wetlands than upstream 

or downstream, the environmental cost may be high. 

Hernandez and Mitsch (2007) found that permanently 

flooded wetlands had lower N2O/N2 ratios of emissions 

than did intermittently flooded wetlands. They also found 

that the ratio was higher in the cold months even though 

the flux rates are much lower then. A full risk assessment 

needs to be made to determine how much of such 

“pollutant swapping” is advisable.

A similar potential exists for Nr mediation in sewage 

treatment. The current practice is to convert ammonia/

ammonium that mineralizes from excreted organic matter 

to nitrate through the nitrification process. As nitratecontaining effluent from sewage treatment flows into 

aquatic systems the nitrate may be denitrified, resulting 

in N2O production if denitrification is not complete. The 

protein consumption by some 301 million humans in the 

U.S. results in the processing of ~ 2 Tg of N annually 

(~18.4 g N/ person/day), much of which flows through 

sewage treatment facilities and ultimately leads to the 

production of between 0.06 and 0.1 Tg of N2O-N /yr in 

aquatic systems or soils to which sewage sludge is applied. 

Tradeoffs among C and N-driven impacts 

Reactive N also contributes to many impacts on the 

environment that are also impacted by other chemical 

species, notably carbon. As depicted in Figure 18, there 

are several points of tangency between the global C and 

N cycles. These are: combustion, agricultural production, 

industrial production, soil and sediment processes, and 

Figure 17: Diagram of the nitrification and 

denitrification processes

Source: Mosier and Parkin, 2007. Reprinted with 

permission; copyright 2007, Taylor & Francis Group LLC 

– Books.
55

end-of-life disposition of products. The implication of these 

interactions is that, in many instances, the perturbation 

of one cycle cannot be fully assessed without including 

effects on the other. For example, proposals to develop 

bio-based products (biofuels, but also other products) as the 

preferable alternative to fossil-based resources are not free 

from impacts. Such trade-offs may involve a single impact 

(e.g., global climate change to which both carbonaceous 

gases and N2O contribute) but may also involve trade-offs 

between impacts that are not easily compared. Figure 19 

shows the latter case in the form of climate change impacts 

(to which C is a principal contributor) versus eutrophication 

impacts (to which nitrogen is a principal contributor) for 

several different biofeedstock-product combinations which 

are evaluated relative to the substituted commercial product 

made from fossil C. A value of 100% on the y-axis would 

mean that the bio-based alternative is no better than the 

fossil-based counter-product, while the negative region 

of the y-axis in Figure 19 represents net C sequestration. 

It is difficult to make direct comparisons across disparate 

impact categories, however Figure 19 suggests that, in 

choosing among alternatives, policies that aim to minimize 

both sets of impacts would be preferred. 

Finding 18: The Committee notes that the effective 

management of Nr in the environment must recognize the 

existence of tradeoffs across impact categories involving 

Nr transformations and the cycling of other elements.

Recommendation 18: The Committee recommends 

that the integrated strategies for Nr management outlined 

in this report be developed in cognizance of the tradeoffs 

associated with reactive nitrogen in the environment 

(consistent with the systems approach of overarching 

recommendations 2 and 3 discussed in Section 6.2 of this 

report). Specific actions should include:

n Establishing a framework for the integrated 

management of carbon and reactive nitrogen;

n Implementing a research program that addresses the 

impacts of tradeoffs associated with management 

strategies for carbon, reactive nitrogen, phosphorus, 

and other contaminants of concern;

n Implementing a research and monitoring program 

aimed at developing an understanding of the combined 

impacts of different nitrogen management strategies 

on the interchange of reactive nitrogen across 

environmental media. 

Figure 18: Combined carbon and nitrogen global cycles 

Source: Miller et al., 2007 (Figure 1, p. 5178). Reprinted with permission; copyright 2007, American Chemical Society. 
56

4.9. Interactions of the N Cascade 

and Climate

Weather and climate vary substantially on many time 

scales including the interannual. Long-term (decadal or 

more) changes in climate as have been predicted by IPCC 

(2007a,b) may have profound effects on the N cycle; 

conversely, changes in the biogeochemical cycle of Nr can 

induce climate forcing. While it is beyond the scope of 

this report to fully address how cycles of C and N interact 

(see Figure 18 for a general treatment of the intersection 

points of C and N cycles), there are several ways in which 

climate impacts the biogeochemical cycle of Nr and vice 

versa (e.g., Yienger and Levy, 1995; Holland et al., 1997; 

Hungate et al., 2003; Hungate et al., 2004; Sutton et al., 

2007; Thornton et al., 2007; Levy et al., 2008; Sokolov 

et al., 2008). These are highly interactive and nonlinear 

systems. The following important interactions are noted:

n Increased deposition of Nr into terrestrial and aquatic 

ecosystems can alter the sequestration of carbon, while 

increased ambient CO2 can change the deposition and 

uptake of Nr. 

n Nitrate flux from fields to surface waters increases 

with increasing rainfall (see Box 5:The Impact of 

Climate Change on Agricultural Discharge of Reactive 

Nitrogen). 

n Increasing temperature can both increase and decrease 

atmospheric loading of particulate matter. 

n Aerosols (PM) have direct and indirect (through cloud 

microphysics) effects on radiative forcing of climate 

and on the hydrological cycle.

n N2O and O3 are greenhouse gases.

n Soil Nr chemistry and emissions of N2O, NH3, and 

NO depend on environmental conditions such as 

temperature and soil moisture.

n The amount of Nr deposited and exported from the U.S. 

depends on meteorological variables including wind 

speeds and convection.

Numerical models, when verified against past climates, 

can provide insight into possible future climates and their 

impacts on the nitrogen cycle. For example, increasing 

temperatures increase the amount of NOx control 

necessary to achieve the same amount of photochemical 

smog control (Bloomer et al., 2009; Jacob and Winner, 

2009). The EPA program for studying the impact of 

climate change on photochemical smog (air pollution 

ozone) production offers a useful model; see Jacob and 

Winner (2009) for an overview. 

Finding 19: The biogeochemical cycle of Nr is 

linked to climate in profound, but nonlinear ways that 

are, at present, difficult to predict. Nevertheless, the 

potential for significant amplification of Nr-related 

impacts is substantial, and should be examined in more 

complete detail.

Recommendation 19: The EPA should support crossdisciplinary and multiagency research on the interactions 

of climate and Nr. To determine the interactions of global 

biogeochemical Nr cycles and climate, the Committee 

suggests that EPA follow a series of steps such as:

1. Select several likely scenarios for global climate from 

the IPCC report for the year 2050.

2. Down-scale statistics or nest regional climate models 

within each of these global scenarios to generate 

meteorological and chemical fields (e.g., temperature, 

relative humidity, winds, precipitation, CO2) for a few 

years around 2050.

3. Run several independent biogeochemical Nr models 

(earth system models that include air/water/land) for 

North America for these years with current Nr and 

emissions and application rates.

4. Rerun models with decreased Nr emissions/application 

to evaluate strategies for controlling impacts such as 

those described in this report.

Figure 19: Comparisons between Global 

Warming Potential (GWP) and eutrophication 

impact categories for various bioproducts 

Abbreviations: BD=Biodiesel; CET=Corn Ethanol; 

CSET=Corn & Stover Ethanol; PLA=Polylactic Acid 

(Corn); RL=Rapeseed Lubricant; SL=Soybean Lubricant; 

STET=Stover ethanol; SWEL=Switchgrass Electricity; 

SWET=Switchgrass Ethanol.

Source: Adapted from Miller et al., 2007 (Figure 2, 

p. 5180). Adapted with permission; copyright 2007, 

American Chemical Society.
57

Box 3: The Impact of Climate Change on Agricultural Discharge of Reactive Nitrogen

The discharge of reactive nitrogen from intensively managed agroecosystems is characterized by a number of 

attributes that often exhibit a high degree of variability: fluctuating material flows associated with the degree of 

nitrogen fixation and the extent of denitrification, the interdependence of crops in rotation, and dependence on 

geography, weather patterns (particularly rainfall intensity, duration, and frequency), soil type, and agricultural 

practices.

One way to gauge the impact of climate change on such systems is to examine the ranges exhibited by historical 

data that collectively encompass the range of impacts that are anticipated. The assumption is that a changing climate 

will systematically alter governing attributes in plausibly predictable ways, for example increased annual rainfall 

and temperature over a large geographic region. The IPCC has provided general climate-induced impacts for world 

regions (IPCC, 2007a,b). 

Based on the simulation model of Miller et al. (2006). Red markers are historical data of discharges according to 

year as reported by Powers (2007). Green bars represent a log-normal distribution.

Figure 20: Probability of given discharge level for nitrate in the watersheds of eastern Iowa

NO3

'89 '96

'88

'94'95

'90

'98 '93 '91 '92

'97

0.000

0.005

0.010

0.015

0.020

0.025

0.030

0.006 0.053 0.083 0.125 0.142 0.172

Discharge, Tg/yr

Probability

Lognormal Distribution

Total NO3 Load from E Iowa

Historical Data

The general impact of climate change on the discharge of reactive nitrogen from agroecosystems can be 

discerned from the information in Figure 20. This figure shows a probability distribution for nitrate discharged 

from the watersheds of eastern Iowa (approximately 50,000 square km), which are dominated by corn-soybean 

agroecosystems (a general description of the region can be found in Kalkhoff et al., 2000). It is derived from 

information on the input of synthetic fertilizers in the region during the period 1989-1999, and includes factors that 

describe the transformation and transfer of Nr once applied. The distribution shown was generated using a Monte 

Carlo technique, details of which can be found in Miller et al. (2006). Also included in Figure 20 (in green) is a 

standard log-normal distribution, which the simulation most closely fits, and independently measured annual nitrate 

runoff data (an output of the system) over the same time period, as reported by Powers (2007). The simulation is 

not perfect, but it does capture the extremes of reactive nitrogen discharge, as represented by data for the years 1993 

and 1998. 

Figure 20 shows that the interannual variation in nitrate discharged is nearly 30-fold during the 11-year 

observation period. While the impact of climate change on such a system cannot be predicted for a given year, 

Figure 20 provides a basis for visualizing shifts in nitrate discharge due to changes in those factors that affect Nr 

transformation and transfer. For example a climate change scenario that predicts a general increase in precipitation 

amount and frequency, other factors being constant, will tend to shift the distribution of Figure 20 to the right, 

resulting in generally higher discharges of nitrate (see for example Vanni et al., 2001; the data point for 1993 in 

Figure 20 corresponds to precipitation in the region that was approximately 1.8 times the long-term annual average). 

Other factors, of course, may amplify or retard such impacts. Understanding whether or not implementation of 

best management practices and advanced technological methods can counteract climate change trends that favor 

increases in discharge would require a series of significant research studies and advances in modeling capabilities. 
58
59

5

Integrated Risk Reduction Strategies 

for Nr

5.1. Importance of Integrated Risk 

Reduction Strategies

Chapters 3 and 4 of this report presented the 

environmental impacts and metrics associated with the 

emission of the various forms of Nr and reviewed ways 

of organizing these into impact “categories.” As noted, 

Nr has many impacts on the environment, impacts 

that are interrelated through the nitrogen cascade. As 

previously stated, the nature of reactive nitrogen demands 

an integrated approach within EPA and across other 

relevant federal agencies, as reactive nitrogen cycles 

through the environment in different forms. A number 

of risk reduction approaches and the importance of 

considering Nr control points in the nitrogen cascade are 

discussed below. 

5.2. Control Strategies for Nr

There are several ways in which the release and control 

of Nr in the environment can be approached. In general 

these can be classified as follows:

1. Improved practices and conservation – in which the 

flux of Nr that creates an impact is lowered through 

better management practices, including those that 

preserve or enhance Nr controlling ecosystem 

services (e.g., on-field agricultural practices, 

controlled combustion conditions, ecosystem function 

preservation and management)

2. Product substitution – in which a product is developed 

or promoted which has a lower dependency on or 

releases less Nr (e.g., N-bearing wastes instead of 

corn grain as a feedstock for biofuels, development of 

alternative power sources such as wind and solar)

3. Transformation – in which one form of nitrogen 

is converted to another form (e.g., nitrification of 

wastewater, denitrification in engineered or natural 

systems)

4. Source limitation – in which the amount of Nr 

introduced into the environment is lowered through 

preventive measures (e.g., controls on NOx

generation)

5. Removal – in which Nr is sequestered from impacting 

a particular resource (e.g., ion exchange)

6. Improved use or reuse efficiency – in which the 

efficiency of production that is dependent on Nr is 

improved (e.g., increased grain yields for lower Nr 

applied), or Nr wasted from one source is reused in 

another (e.g., algal farming)

Effective management of Nr requires combinations 

of these approaches; none is a perfect alternative for 

controlling Nr in the environment. Table 14 provides 

a summary of the pros and cons of each of these 

approaches.

5.3. Management Strategies for Nr in the 

Environment

Four types of management strategies for the control of 

Nr, and other pollutants, in the environment have evolved 

over the past 40 years:

1. Command-and-control – in which an entity’s 

discharge of pollutants is regulated through a series 

of permitted limitations on emissions, violations of 

which may result in penalties being assessed

2. Government-based programs for effecting a policy, 

such as directed taxes, price supports for a given 

commodity, subsidies to bring about a particular end, 

and grants for capital expansion or improvement

3. Market-based instruments for pollution control in 

which market trading schemes are used to bring about 

a desired policy end, often at reduced overall cost.

4. Voluntary programs in which desired ends are 

achieved using private or government-initiated 

agreements or through outreach and education.

5.3.1. Command-and-control19

Policy makers have traditionally used command-andcontrol strategies requiring individuals and dischargers 

to meet mandatory guidelines. Such an approach evolved 

as the country was gearing up to meet the requirements 

first established nationally through the CWA and CAA 

enabling legislation in the 1970s. Because U.S. capabilities 

to monitor contaminant concentrations and predict 

environmental impacts were, generally, rudimentary, early 

emphasis was placed on “technology-based” approaches 

for managing emissions. This resulted in the promulgation 

of “best practicable technology” controls, and eventually 

“best available technology” controls, the idea being that 

mandating some level of control, even with uncertain 

improvements on impacts, would be better, and less 

arbitrary, than other approaches of the time. 

19 Based on Models in Environmental Regulatory Decision-Making, National Research Council, 2007.
60

Nevertheless, both the CWA and the CAA had 

more specific goals that were aimed at protecting 

human health, public welfare, and ecosystem health. 

For example, the CAA required states to develop 

implementation plans (SIPs), the approval of which 

depended on their ability, once implemented, to meet 

ambient clean air standards. Likewise, the CWA 

required greater controls to be implemented for certain 

water bodies for which technology-based limits alone 

were insufficient to meet standards (this became the 

TMDL program). 

Over time, and as our abilities to monitor, predict, and 

understand impacts improved, it became possible, or at 

least plausible, to tailor emission levels on a source-bysource basis, allowing the firm in question to decide its 

own technological approach. Thus permits, which place 

strict limits on the amount of pollution a firm is allowed 

to discharge over a specified period of time, have become 

the main method for managing the majority of point 

source contaminants, including the various forms of Nr, 

in the environment. 

While the CWA has had considerable success in 

controlling point source discharges, it has been largely 

unsuccessful in limiting nonpoint discharges, and it 

is these sources that are particularly important for 

managing nutrient flows into receiving waters. The 

National Research Council has addressed this deficiency 

and pointed to the need to fully implement TMDL plans 

and establish numerical nutrient standards for nutrients 

(National Research Council, 2008b, 2009).

5.3.2. Government Taxes and Subsidies to 

Achieve Policy Ends

Government taxes and subsidies have created a 

variety of results, some in conflict with and some to 

further the ends of Nr management. Examples include 

U.S. agricultural and land-use policies, energy and 

transportation policies, and both point and nonpoint 

source mandated controls on N-bearing aquatic resources, 

including domestic and industrial wastewaters and 

agricultural runoff. 

Current and future energy policy with respect to 

vehicle efficiency and biofuels will help determine 

the amount of Nr released into the environment from 

these sources. Some states have chosen to place modest 

taxes on fertilizer containing Nr, though the demand 

impact is slight at best. However, revenues may be 

dedicated to improved Nr utilization efficiency. Crop 

subsidies and crop insurance may at times expand land 

use and even encourage increased use of fertilizers, 

effectively increasing Nr in the environment. There 

are various agricultural conservation programs in 

the U.S. administered by the USDA. These include 

the Conservation Reserve Program and the Wetland 

Reserve Program (CRP and WRP). The former takes less 

suitable land out of cultivation and the latter encourages 

wetland protection and restoration. Both can contribute 

to better Nr management. The Environmental Quality 

Incentives Program (EQIP) directly subsidizes nutrient 

management efforts by crop and livestock producers. 

Of concern to the Committee is the need for more 

Table 14 : Advantages and limitations of various approaches to Nr control in forestry and agriculture

Control strategy Advantages Limitations

Improved practices, 

conservation

Lessens one or more impacts; 

utilization of existing ecosystem 

services

Education cost; availability and cost 

of preserved lands

Product substitution Lessens the need for Nr, allows for 

more targeted uses of Nr

Questions of acceptability, 

technological issues

Transformation

Reduces one or more impacts 

to which Nr contributes, for 

denitrification closes the 

nitrogen cycle; utilizes natural 

biogeochemical processes that may 

be available ecosystem services

May contribute to other impacts; human presence has modified and diminished ecosystem service values

Source limitation

Reduces one or more impacts to 

which Nr contributes

Decreased crop yields, in some 

cases few viable alternatives yet 

developed

Removal

Reduces one or more impacts to 

which Nr contributes; natural land 

features/processes and ecosystem 

services may be used

Residuals containing Nr must still 

be managed effectively; availability, 

location and cost of land for natural 

or enhanced Nr removal

Improved efficiency Reduces the need for Nr Research and education costs
61

effective approaches aimed at encouraging farmers 

and land managers to adopt proven conservation and 

Nr management practices in fields and feedlots. The 

extent of proven practices, such as variable rate fertilizer 

application and installation of stream buffers, fall far 

below today’s technological frontier.

5.3.3. Market-Based Instruments for Pollution 

Control20

A fundamental shift in environmental management 

philosophy was initiated with the 1990 Clean Air Act 

Amendments, which combined regulatory requirements 

with market flexibility allowing lower compliance costs 

through tradable credits. Most market-based policy 

instruments operate on the principle that if the regulatory 

framework or some other factor sufficiently alters 

the relative value of available decision choices for an 

individual or firm, subsequent decisions they make will be 

in alignment with the policy makers’ objective. 

As an example, if a government wants to limit 

pollution in a river where a number of polluters 

discharge, it need not adopt a uniform command and 

control limit on each discharger. Instead, a regulatory 

cap on the total pollutant loadings can be established and 

individual permit limits can be issued to all dischargers, 

with provisions that allow the dischargers to trade 

between their individual limits as long as the overall cap 

is not exceeded. Those dischargers having low pollution 

control costs will have incentive to control more 

pollution than their permit limit and thus generate water 

quality credits that can be sold to dischargers with high 

costs of pollution control. Because the overall cap on the 

pollutant is fixed, the regulatory goal is achieved. Water 

quality trading thus brings about the desired reduction in 

pollution level at lower cost than if all dischargers were 

required to use traditional onsite treatment technology. 

Water quality trading also encourages cost-effective 

pollution control investment by giving each firm a clear 

economic signal to invest in new technology to reduce 

pollution at a level that corresponds to the market value 

of the permit.

As with control strategies for Nr, there is no one 

universal market-based strategy that is applicable to 

every policy maker’s objective. For example, the nature 

of incentives available to and effective with producers 

involved in over-fishing is different from landowners 

providing environmental amenities. In the former case, 

the objective is to restrict the intensity of fishing. In 

the latter case the objective is to encourage private 

landowners to provide environmental goods and services 

at the lowest cost possible.

Evolution of new market-based strategies is a 

continuous process. Most strategies have been customized 

over time to meet local needs. One can group such market 

based approaches under the following conceptual headings:

1. Water Quality Tradable Credits – Every polluting 

entity is allowed to discharge pollutants up to a certain 

pre-determined limit, defined in concordance with 

the terms of the CWA. The entities discharging less 

than their allocated limit generate credits. Under this 

strategy, credits can be traded with other polluting 

entities that have exceeded their allocated limit 

provided that water quality standards are not exceeded. 

2. Auction-Based Contracting – Environmental or 

conservation contracts are auctioned, where individual 

landowners place their bids to provide such goods or 

services from their land. Two factors jointly determine 

the selection of the bids: the amount of the bid and the 

expected value of the environmental or conservation 

benefit resulting from accepting the bid.

3. Individual Transferable Quotas – An individual 

transferable quota (ITQ) is an allocation privilege 

to extract a specified quantity of a resource among 

a selected number of quota holders. The distinctive 

feature of the ITQ is that the privilege is transferable 

or leasable. An ITQ may be a right to produce under 

favorable circumstances, such as a tobacco quota 

when tobacco production would normally be limited.

4. Risk Indemnification for Specified Behavior – An 

example of this is crop insurance designed to protect 

farmers from uncertainty in the adoption of best 

management practices that provide a public good but 

are inherently riskier.

5. Easements – Conservation easements or conservation 

servitudes refer to the case in which a landowner 

enters into a legally binding agreement to surrender 

certain property rights for a specified period of 

time, either voluntarily or for compensation. Such 

arrangements usually provide public goods relative to 

the environment or conservation.

Policy maker objective, local conditions, and several 

other factors determine the suitability of a particular 

market based strategy. For example, water quality 

trading is well suited where there are a variety of 

dischargers at different levels of contribution and with 

varying control costs. A policy framework that facilitates 

the emergence of multiple options for dischargers to 

meet their permit limits, such as buying from more 

efficient controllers of discharge or investing in new 

equipment to achieve further reductions, is likely to 

accomplish the desired level of water quality at the least 

possible cost to the economy. Table 15 illustrates the 

potential effective application of a number of marketbased approaches in specific situations. Accompanying 

this chapter are two examples of the application of 

20 Based on Canchi, D., P. Bala and O. Doering, Market Based Policy Instruments in Natural Resource Conservation, Report for the Resource 

Economics and Social Sciences Division, NRCS, USDA, Washington D.C., March 3, 2006, pp. 4-9.
62

market-based approaches for the design of water 

quality trading schemes for Nr in watersheds (Box 4: 

Water Quality Trading to Meet the Long Island Sound 

Wasteload Allocation in Connecticut, and Appendix C: 

Water Quality Trading in the Illinois River Basin). 

Table 15 shows pair-wise comparison between 

different market-based strategies. The objective and 

the incentive structure of the participants determine the 

suitability of one market based strategy over another. 

Each pair of cells briefly lists the most relevant set of 

conditions for which the respective strategy may be 

optimal (left cell points to strategy at the top of the 

column and right cell points to the strategy at the end 

of the row). Consider the two strategies (illustrated 

below): Auction-Based Contracting and Tradable 

Credit. If the participation of every private entity is 

essential, then Auction-Based Contracting works best. 

For example, if the objective is to preserve a large tract 

of privately owned contiguous land, Auction-Based 

Contracting is the appropriate strategy. This requires 

the participation of every private land owner to set 

aside a portion of their land. An auction designed to 

reveal the individual’s land owner’s reserve price for 

participation leads to the most efficient solution. A 

classic example of this is the Australian Government 

of Victoria’s auction based Bush Tender program. Here 

it was essential to enlist blocks of land with particular 

hydrological and other characteristics to maximize the 

reduction in salinity and provide other environmental 

benefits. (Department of Sustainability and 

Environment, 2008). The goal was to bring about the 

reintroduction of native vegetation, its protection, and 

management where it would slow the development of 

salinity in soils. Offsets would not accomplish this, and 

as one looks at Table 15’s characteristics of alternative 

approaches, neither would quotas, or insurance for 

BMPs. Easements might be used, but auctions were 

much more cost effective and more suited to the long 

term management commitment needed, as indicated 

in the second row of Table 15 under “Easements.” 

Compared to this, if the objective is an overall 

reduction of a pollutant regardless of the individual 

private entity’s contribution to the abatement, the 

Tradable Credit strategy with a cap is more appropriate. 

As another example, if aggregate depletion is of 

concern (as with fisheries) then individual transferable 

quotas are appropriate. However, auction-based 

contracting is preferable to individual quotas when no 

offsets exist.

Although there are significant differences between 

water and air quality trading, there are also several 

potential barriers to effective trading systems for 

both media. These are related to: accountability and 

monitoring; establishing standards and management 

goals; complexities of cross media and multiple source 

trading, including parity of sources; insurance that 

outcomes would reduce risk (environmental benefit); 

economics and marketability of traded credits; and 

transparency of the program, including public outreach 

and stakeholder involvement.

5.3.4. Biophysical and Technical Controls (control points) on Transfer and Transformations of 

Nr in and between Environmental Systems

Within the nitrogen cascade there are a number of 

places where the flow of Nr is constrained or regulated, 

either by nature or by human intervention, or a 

combination of the two. This report refers to these places 

in the cascade as “control” points. The control points 

may restrict the flow of Nr species within environmental 

systems (atmospheric, terrestrial, aquatic) or between 

them. The control points vary from primary controls 

where Nr is minimized through conservation measures or 

through after-the-fact measures that attempt to convert Nr 

that is emitted or not fully used to nonpolluting products 

(such as conversion to N2 by denitrification or through 

long-term storage). The discussion of control points in 

this section is primarily focused on biophysical controls in 

terrestrial and aquatic environmental systems. However, 

the section concludes with a discussion of possibilities for 

decreasing NOx emissions from combustion.
63

Table 15: Summary of market-based instruments for pollution control with conceptual examples

Auction Based 

Contracting

Individual 

Transferable 

Quotas

Insurance for the 

Adoption of BMPs Easements

When there 

exist no 

offsets; The 

participation 

of every 

private entity 

is critical;

Offsets are 

possible; 

Aggregate 

effect is of 

concern, 

not each 

individual 

entity’s 

contribution;

When the 

depletion is of 

concern;

When the 

discharge is of 

concern;

Homogenous 

polluters; 

Offsets not 

feasible; 

Excessive 

pollution is 

primarily 

to mitigate 

uncertain 

profits; 

Modest 

short-term 

objective;

Not 

homogenous 

polluters; 

Offsets are 

possible; 

Pollution is 

an absolute 

consequence 

of the 

production 

process;

Unidirectional; 

When offsets 

are not 

possible; One 

entity retiring 

more property 

rights cannot 

trade with the 

other retiring 

less property 

rights.

Bidirectional; 

Offsets are 

possible; 

Requires 

specific action 

on the part of 

the participant 

to accomplish 

the objective;

Tradable 

Credit

Aggregate 

depletion is of 

concern;

When there 

exist no 

offsets; The 

participation of 

every private 

entity is 

critical;

Tied to a 

production 

process; 

When risk 

averseness 

of the entity 

can be used 

to motivate 

participation;

Not tied to any 

production 

process; 

Suited for 

motivating 

participants 

to engage in 

secondary 

activities;

Auction based 

contracting 

can be seen 

as a refined 

and improved 

cost-efficient 

alternative to 

easements;

Designing of 

auction based 

contracting 

requires 

considerable 

professional 

expertise; 

Auction 

Based 

Contracting

Discharge of 

effluents is of 

concern;

Depletion of a 

resource is of 

concern;

Retirement 

of rights is of 

concern;

Acquisition 

of rights is of 

concern;

Individual 

Transferable 

Quotas

No uncertainty; 

No action 

required on 

the part of the 

participant;

Tied to a 

production 

process:

Insurance 

for the 

Adoption of 

BMPs

Each pair of cells briefly lists the most relevant set of conditions for which the respective strategy may be optimal (left 

cell points to strategy at the top of the column and right cell points to the strategy at the end of the row).

Box 4: Water Quality Trading to Meet the Long Island Sound Wasteload Allocation in Connecticut

Pollutant trading is increasingly being promoted as a cost-effective means for attaining water quality standards. 

Connecticut and New York have been working with the EPA Long Island Sound Study (LISS) for more than 

20 years to address low oxygen conditions (hypoxia) in Long Island Sound that have been linked to excessive 

loadings of nitrogen. A Total Maximum Daily Load (TMDL) for nitrogen, drafted by the two states and approved 

by the EPA in 2001, set a 58.5% nitrogen reduction target in 2014 from point and nonpoint source/stormwater 

sources. Connecticut has initiated a point source trading program for 79 municipal sewage treatment plants (STPs) 

to facilitate implementation of the TMDL wasteload allocation (WLA) and is investigating the potential for 

incorporating nonpoint source/stormwater into the existing Nitrogen Credit Exchange (NCE).

Several prerequisite conditions essential to the success of the current point source trading program have been 

met. Briefly, (1) all the STPs contribute to the same water quality problem; (2) the technology to remove N and meet 

the targets exists; (3) there are compelling member benefits to participate, especially cost savings; (4) sources can 

easily be monitored and tracked by end-of-pipe monitoring; (5) credit cost calculations are based on established and 

agreed upon protocols founded in state legislation; (6) sources of N are diverse and create viable supply and demand 

conditions while reducing overall cost, with close control by a Nitrogen Credit Advisory Board (NCAB); and (7) 

transaction costs are low relative to credit prices. In operation since 2002, the NCE has proven to be a viable and 

effective mechanism for meeting the nitrogen WLA.

The economic record of the NCE demonstrates the vigor of trading over the first five years of completed trades 

from 2002 to 2006 (Table 16). In sum, more than 10 million credits have been traded on the NCE, representing more 

than $22 million in economic activity.
64

The use of geographically-based trading ratios is instrumental to the relative cost of meeting N reduction limits at 

the 79 treatment plants, which are scattered throughout the state (Figure 21). Because N is reactive as it travels down 

rivers into the Sound, and the Sound’s currents further affect relative impacts as they transport N and the resulting 

algal blooms to the hypoxic areas at varying efficiencies, location of each treatment plant makes a difference in 

relative impact on dissolved oxygen per pound of N discharged at end of pipe. Generally, the closer a POTW is to 

the edge of the Sound, and the closer to the hypoxic zone, the higher the trading ratio (Figure 22). For plants with 

high trading ratios, economics often favor treatment, while those with lower ratios may find the purchase of credits 

economically advantageous over treatment.

The point source NCE does not reflect a free market approach to trading. Demand is set by the annual general 

permit limit and supply of credits is constrained by the availability of Clean Water Fund dollars and the timing and 

location of N removal projects. Credits are bought and sold from the state, thus the number of credits purchased does 

not need to match the number of credits sold (as would typically be true in a tradable permit system). Nevertheless, 

there is a tendency towards implementing cost effective projects as sewage treatment plant authorities decide 

whether it is less expensive to treat or buy credits, and try to predict when that break-even point might occur that 

would warrant application for project funding.

Trading

Year

Credit 

Price

(Dollars)

Purchased 

(Dollars)

Sold

(Dollars)

Purchased

(1000 

Credits)

Sold (1000 

Credits)

2002 $1.65 $1,317,223 $2,357,323 798 1,429

2003 $2.14 $2,116,875 $2,428,636 989 1,135

2004 $1.90 $1,786,736 $2,659,804 940 1,400

2005 $2.11 $2,467,757 $1,315,392 1,170 623

2006 $3.40 $3,828,114 $2,394,956 1,126 704

Total $11,516,705 $11,156,111 5,023 5,291

Table 16: Performance of the Nitrogen Credit Exchange

Source: Connecticut Department of Environmental Protection, 2007

Figure 21: Relative nitrogen discharge (lbs/day) from 79 POTWs.

Source: Connecticut Department of Environmental Protection, 2007
65

Incorporating a nonpoint source/stormwater (NPS/SW) component into the existing point source trading program 

presents some difficult challenges. Among the seven prerequisite conditions listed above that are well met by the 

current point source program, NPS/SW trading does not provide compelling economic benefits for members; NPS/

SW N is difficult to quantify and track; credit cost estimation does not have a strong foundation in any existing 

programs; NPS/SW credit costs, though geographically diverse, may not result in significant implementation 

savings; and transaction costs (or time spent negotiating the ground rules for NPS/SW trading) may be considerably 

higher than for point source credits. Many of these obstacles can be overcome by deferring to models and textbook 

costs and efficiencies for NPS/SW BMPs. Tracking will still be a challenge because of the sheer number and 

distribution of BMPs that can be applied throughout the state.

Basic economic principles suggest that a free-market arrangement will not produce many NPS/SW credits for 

market. Costs are much higher than for point source credits and a regulatory approach must therefore be instituted to 

formalize the load allocation for nitrogen and to structure participation by municipalities. 

If a NPS/SW trading component were to be added in the future, it would most likely also be an incentive-based 

program rather than a free-market approach. Nitrogen is difficult and costly to control in Connecticut’s urban/

suburban setting, and reductions are unlikely to be cost- competitive with POTW credits in a free market system. 

However, because municipalities are required to implement the Phase II stormwater permit, and various federal, 

state, and local programs require or emphasize NPS/SW management, there may be benefits for an incentive-based 

approach to offset some of those costs. For example, payment for NPS/SW reductions at the same credit prices paid 

to POTWs under the NCE would help defray costs and encourage additional nitrogen reductions from stormwater/

NPS sources. Connecticut and the NCAB will continue to evaluate and explore the viability of these options.

Market approaches and trading can lower costs and increase economic efficiency of Nr control. The approaches 

may well have to be situation specific and depend on a structured regulatory framework to create the market or 

trading opportunity. As with the 1990 Amendments to the Clean Air Act, the design of market based instruments 

is a product of technical capability, regulatory design, and public preference. Implementation can be tedious but 

the benefits in efficiency substantial, even after being balanced against equity concerns. However, there can be 

something of a geographical and supply/demand mismatch between nonpoint sources and point sources that might 

be trading partners. Ribaudo and Nickerson identify only 142 of 710 eight-digit Hydrologic Watershed Units 

containing waters impaired by nitrogen where trading would be most likely (Ribaudo and Nickerson, 2009).

Further, the cost for management will be enormous. EPA’s Clean Water Needs Survey (EPA, 2008b) has identified 

more than $200 billion in wastewater management infrastructure needs, and those needs do not fully address nutrient 

control from both traditional point and nonpoint/stormwater sources or consider alternative technologies.

Figure 22: Trading ratios for municipalities in Connecticut.

Source: Connecticut Department of Environmental Protection, 2007
66

Biophysical controls in terrestrial environmental 

systems

As indicated in Figure 2, approximately 36 Tg of 

new Nr is introduced into the U.S. each year. This new 

Nr is derived from sources that include consumption of 

~11 Tg of synthetic N fertilizer, ~8 Tg of N that is fixed 

biologically by crops, and ~ 5 Tg that is emitted from 

fossil fuel combustion annually. This N is used to produce 

food and fiber (~15 Tg) or is formed during electrical 

generation, industrial production, or transportation. 

Efforts to decrease the creation of new Nr should first 

look to conservation. 

Reduction in use of fossil fuel and/or decreased Nr 

emission can come through a variety of mechanisms such 

as more energy-efficient industrial processes, homes, and 

vehicles. Further gains are possible through conservation 

practices and alternatives to wasteful approaches, such 

as improving public transportation to minimize use of 

personal automobiles, and use of local products that do 

not require long-distance shipping. 

Improvements in food and fiber production and 

changes in diet can also play an important role in limiting 

Nr. Because agriculture is the largest consumer and 

producer of Nr, consumption of fertilizer N could be 

decreased by changes in diet and increasing fertilizer N 

use efficiency in crop and fiber production systems. The 

control points discussed in this section include: protein 

consumption in the human diet; removing croplands that 

are highly susceptible to Nr loss from crop production; 

decreasing fertilizer N demand by increasing fertilizer 

use efficiency in crop and fiber production, as well as 

on residential and recreational turf grass; and better 

management of Nr in manure from livestock production 

in CAFOS.

Decreasing the amount of fertilizer N needed 

through changes in human diet

Along with increasing fertilizer N use, continued high 

intake of protein in developed countries and changes in the 

diet of people in developing countries will likely lead to 

greater N losses from global food production in the future. 

The first aspect of changes in food production concerns 

the increasing protein consumption that is occurring as 

global population increases and gets wealthier. This is 

likely to require increased N input into food production 

(Naylor et al., 2005; Galloway et al., 2007). 

The average protein supply per person in developed 

countries is presently ~100 g per day, while in the 

developing countries it is only ~65 g per day (Food and 

Agricultural Organization Statistical Database (FAO 

FAOSTAT, 2010a). There is a direct proportionality 

between protein and nitrogen composition of food (ca 

0.16 g N per 1 g protein). On average in 1995, developed 

countries consumed ~55% of total protein from animal 

sources while developing countries derived ~25% of total 

protein from animals. Protein consumption was highest 

in the U.S. and Western Europe, ~70 and ~60 g animal 

protein per person per day, respectively. In 2003, total 

protein consumption in the U.S. was 115 g per person 

per day (74 derived from animals and 41 from vegetable) 

(FAO FAOSTAT, 2010a). In developing countries, 

the greatest change in animal protein consumption 

has occurred in China where the consumption of meat 

products has increased 3.2 fold (from ~10 to ~32 g per 

person per day) since 1980. In Sub-Saharan Africa there 

has been no increase in either total (~50 g per person 

per day) or animal protein (~10 g per person per day) 

consumption during the past 30+ years (Mosier et al., 

2002). 

The reason for focusing on the consumption of 

animal protein is that more N is needed to produce a 

unit of animal protein than an equal amount of grain 

protein. Bleken et al. (2005) note that the N cost of 

animal production in Norway and the Netherlands was 

approximately five units of N in feeds for each unit of 

N produced. Approximately 2.5 units of N are required 

to produce a unit of wheat protein-N. Bequette et al. 

(2003) report that dairy cattle consume four units of N 

in feeds (including forage and grains) for every unit of 

N that appears in milk. Using a range of efficiencies 

for animal production practices, Kohn et al. (1997) 

estimated that 4 to 11 units of fertilizer N would be 

used in a whole farm system to produce a unit of milk 

protein. This ratio would be lower when using legume 

N to feed cattle, as is commonly done. Based upon the 

extra N required to produce animal protein compared 

to grains, continued high protein consumption in 

developed countries and changes to higher protein diets 

in developing countries will likely increase N input and 

losses in food production.

Moderating this increase by decreasing the average 

amount of total protein consumed in developed countries 

is one mechanism of limiting part of the expected 

increased N requirement in food production. One 

example of a country with a healthy diet and moderate 

consumption of animal protein is Italy in 1963. At that 

time, food supply was adequate to ensure sufficient 

nutrition to all groups of society (Bleken, 1997). Total 

protein consumption was 85 g per person per day, and 

consumption of animal protein was 32 g, roughly half 

of the current U.S. diet, and yet much higher than the 

average of developing countries. Another example is 

Japan, where animal protein consumption has traditionally 

been low, although it has increased from 25 g in 1963 to 

54 g animal protein per person per day in 1995. In the 

same period the total protein consumption has increased 

from 73 g to 96 per person per day.

Bleken (1997) analyzed the relation between human 

diet and global N need for food production. Her analysis 

indicates that the total N needed for diets with high 

animal protein intake (comparable to many industrialized 

countries today) is almost twice as high as the N needed 
67

for the average diet in Italy 1963, or for Turkey in 1993. 

Based on her analysis, the Committee assumes that in 

the high-N input regions, per capita N need for food 

production may be reduced by 45%, which would reduce 

present-day N inputs by 15% worldwide. 

Switching to a lower protein diet may not, however, 

reduce N losses if the new diet includes increased 

quantities of fruits, vegetables, and nuts, in addition to 

staple grains, beans and pulses. Vegetables, fruit, and 

nuts are high value crops that typically require large 

inputs of fertilizers and pesticides when produced at a 

large, commercial scale, and N fertilizer losses can be 

considerably larger than for grain crops. Having a very 

diverse diet that includes a wide range of high-value 

fruits and vegetables available year round (whether 

they are in-season locally or not) also has consequences 

for N inputs/outputs from agriculture – both within the 

U.S. and globally. EPA and USDA are encouraged to 

develop programs that stress how both human health and 

environmental health will improve with a greater focus on 

the human diet. It has been estimated that 30% – 40% of 

the food prepared for consumption in the U.S. is wasted 

(Kantor et al., 1997; Hall et al., 2009). Thus, additional Nr 

may be conserved by decreasing the amount of food that 

is wasted. 

Removing croplands that are susceptible to Nr 

loss from crop production

An analysis of NO3

-

 loading in the Mississippi 

River Basin (Booth and Campbell, 2007) provides 

estimates of N input from agricultural lands. Similar 

estimates were provided by Del Grosso et al. (2006). 

Recommendations in this analysis are essentially the 

same as those arrived at in the original national hypoxia 

assessment, which suggested that the most leaky lands 

be taken out of production (Doering et al. 1999). Booth 

and Campbell state:

Nitrogen derived from fertilizer runoff in the 

Mississippi River Basin (MRB) is acknowledged as 

a primary cause of hypoxia in the Gulf of Mexico. 

To identify the location and magnitude of nitrate 

runoff hotspots, and thus determine where increased 

conservation efforts may best improve water quality, 

we modeled the relationship between nitrogen inputs 

and spring nitrate loading in watersheds of the MRB. 

Fertilizer runoff was found to account for 59% of 

loading, atmospheric nitrate deposition for 17%, 

animal waste for 13%, and municipal waste for 

11%. A nonlinear relationship between nitrate flux 

and fertilizer N inputs leads the model to identify a 

small but intensively cropped portion of the MRB 

as responsible for most agricultural nitrate runoff. 

Watersheds of the MRB with the highest rates of 

fertilizer runoff had the lowest amount of land enrolled 

in federal conservation programs. Our analysis 

suggests that scaling conservation effort in proportion 

to fertilizer use intensity could reduce agricultural 

nitrogen inputs to the Gulf of Mexico, and that the 

cost of doing so would be well within historic levels 

of federal funding for agriculture. Under this simple 

scenario, land enrolled in conservation programs 

would be increased by about 2.71 million hectares, a 

29% increase over 2003 enrollments, while land taken 

out of traditional fertilized agriculture and enrolled in 

conservation programs would constitute about 3% of 

2003 fertilized hectares.

The Booth and Campbell approach places the leakiest 

intensively cropped lands into government programs 

like the Conservation Reserve Program – where they 

would be put into grass or cover crops. Doering et al. 

(1999) had a somewhat different approach. Under their 

analysis, nitrogen use or nitrogen loss reductions were 

imposed on agriculture, and the U.S. Agricultural Sector 

Mathematical Programming (USMP) model adjusted 

crop rotations, tillage practices and fertilizer inputs 

within the Mississippi Basin – meeting the given Nr 

constraint while maximizing producer and consumer 

welfare. The model favored those crops and cropping 

systems at different points in the landscape having low 

nitrogen leakage. Where the model could not find a crop 

production system having positive returns while meeting 

the Nr restrictions, the land was retired from production. 

This analysis suggests opportunities for maintaining land 

in agricultural production while still reducing Nr losses 

through better matching of land characteristics with crops 

and cropping systems.

This 1999 analysis of the Mississippi Basin was 

carried out in the context of cost effective approaches – 

starting with the most cost-effective (in terms of producer 

and consumer welfare) and moving to less cost-effective 

approaches as more and more nutrients were controlled. 

This included both restriction of fertilizer inputs, buffers, 

and wetland remediation as well as the land use changes 

and crop rotations referred to above. The suggestions 

presented by the Committee for Nr reductions that could 

be achieved from agriculture with existing technology 

are consistent with the cost effective approaches in the 

1999 Hypoxia Assessment’s economic analysis. Cost 

effectiveness and alternative cropping systems were 

considered in the SAB report, Hypoxia in the Northern 

Gulf of Mexico: An Update by the EPA Science Advisory 

Board (U.S. EPA SAB, 2007) but unfortunately as 

pieces from individual study examples rather than as an 

integrated approach like the 1999 Hypoxia Assessment 

(Doering et al., 1999).

Decreasing fertilizer N demand by increasing 

fertilizer use efficiency in crop and fiber production

The largest input of Nr in North America is N 

fertilizer used for crop production. The mean annual 

N fertilizer input to North America between 1999 and 

2003 was 12.5 Tg. Of this fertilizer N, 66% was used to 

fertilize cereal crops, mainly corn and wheat (Dobermann 

and Cassman, 2005). 
68

As previously discussed, corn yield in the U.S. has 

increased (from an average of 100 bushels per acre in 

1985 to 136 bushels per acre in 2005) as a result of 

improved nutrient and pest management, expansion 

of irrigated area, conservation tillage, soil testing, and 

improved crop genetics (yield and pest resistance) 

(Council for Agricultural Science and Technology 

[CAST], 2006). From 1980 to 2000, N-fertilizer use 

efficiency (NFUE, kg grain produced per kg applied N, 

or kg grain / kg N) increased from 42 to 57 kg grain / kg 

N, a 35% efficiency gain during a period when average 

U.S. corn yields increased by 40% (Fixen and West, 

2002). Despite this steady increase in NFUE, the average 

N fertilizer uptake efficiency for corn in the north-central 

U.S. was 37% of applied N in 2000 based on direct 

field measurements (Cassman et al. 2002). These results 

indicate that greater than 50% of applied N fertilizer 

is vulnerable to loss pathways such as volatilization, 

denitrification, runoff, and leaching. The results also 

suggest there is substantial room for improvement in 

N efficiency currently achieved by farmers. Although 

progress has been made to increase both cereal yield 

and NFUE, a concerted effort to further increase NFUE 

remains a logical control point to reduce production 

costs, because N fertilizer represents a significant input 

cost, and to limit Nr leakage (e.g., NH3, NOx, N2O, 

NO3

-

) from agroecosystems. 

The goal of reducing Nr while sustaining adequate 

rates of gain in cereal production to meet expected 

food demand will require increases in NFUE, which in 

turn will require innovative crop and soil management 

practices. This need is exacerbated by the recent increase 

in demand for corn to produce ethanol biofuel. The 

concept of improved N synchrony (practices that better 

match the amount, timing, and geospatial location of 

applied N to crop-N demand and the N supply from 

indigenous soil resources) is generally viewed as the 

most appropriate approach for improving NFUE (e.g., 

Appel, 1994; Cassman et al., 2002). The challenge is to 

attain greater synchrony between crop N demand and the 

N supply from all sources (e.g., soil, fertilizer, organic 

inputs such as manure, compost, or green manures) 

throughout the growing season. Losses from all N-loss 

mechanisms increase in proportion to the amount of 

available N present in the soil profile at any given time.

Several promising technologies and combinations of 

technologies have emerged in recent years. Significant 

increases in NFUE are often achieved through reducing 

N fertilizer use by 10 to 30%, while still maintaining 

or even slightly increasing yields (Giller et al. 2004). 

Figure 23 indicates where the greatest gains in NFUE 

are expected to be realized from investments in different 

technology options. Improvements in crop and soil 

management practices will contribute to higher NFUE 

by achieving greater congruence in timing of the supply 

of applied N with crop-N demand and the N supply from 

indigenous soil resources. While there is relatively small 

scope for specific biotechnology traits to improve NFUE, 

overall improvement in crop genetics from commercial 

breeding efforts that focus on increasing yield and 

yield stability will continue to play a significant role in 

improving overall NFUE. However, large investments 

in research, extension education, and technology 

transfer will be required, and significant incentives must 

be implemented, to achieve the degree of improved 

synchrony needed to make substantial improvements 

in NFUE. The need to accelerate the rate of gain in 

crop yields to meet increasing demand for human food, 

livestock feed, and biofuels represents an additional 

new challenge. Crop prices are expected to rise as they 

more closely track the price of petroleum (Council for 

Agricultural Science and Technology, 2006). Higher crop 

prices will motivate farmers to achieve higher yields, and 

higher crop yields require a greater amount of N uptake 

to support increased biomass production (Greenwood et 

al., 1990). Therefore, an explicit emphasis on developing 

technologies that contribute to both increasing yields 

and NFUE will be needed to ensure that the goals of 

food security, biofuel production, and protection of 

environmental quality are met.

Alternatives to current urban landscaping practices

Section 2.2.4 discussed the use of turf grasses as a 

prominent feature in U.S. urban landscapes with over 

1 TgN used to fertilize lawns each year (Table 9). 

New developments are most amenable to landscaping 

practices that may minimize the need to use supplemental 

fertilizer. These practices include preservation of the 

natural soil profile, use of turf types that require little 

or no fertilizer, minimizing turf areas, using organic 

Figure 23: The likely impact of research 

investment in increasing N fertilizer use 

efficiency

Source: Giller et al., 2004 (Figure 3.2, p. 48). Reprinted 

with permission from Island Press; copyright 2004, 

SCOPE.
69

maintenance techniques, and choosing alternatives 

to lawns and exotic plant species such as naturalistic 

landscaping. Many of these practices are part of a low 

impact development philosophy, which can also combine 

other best management practices to mitigate the effects 

of impervious cover and landscape changes. Existing 

development is also amenable to many of these practices, 

especially conversion of typical residential 

and commercial lawns to natural landscapes and 

retrofitting other BMPs that promote infiltration, such as 

rain gardens.

Structural and non-structural Best Management 

Practices (BMP) to treat runoff

There are probably hundreds, if not thousands, of 

BMPs that have been designed and manufactured to treat 

runoff from both urban and agricultural lands. Whether 

applied to new development or existing agricultural or 

urban land use, most follow basic principles that simulate 

natural land features and processes that remove pollutants 

from runoff. They promote infiltration to take advantage 

of the cleansing value of passage through soils and to 

reduce runoff volumes, and provide for biological or 

chemical conditions that help remove pollutants (NRC, 

2008b, 2009). 

The most notable of the processes for managing Nr is 

providing conditions that are adequate to denitrify Nr in 

the waste-stream in a process called biological nitrogen 

removal (BNR). BNR simply creates conditions that 

convert initial forms of nitrogen to nitrate via oxidation, 

and convert nitrate to dinitrogen gas by providing 

conditions (especially high carbon and low oxygen) 

where the denitrification process can occur. These 

simulate natural conditions such as nitrification that 

occurs in oxic soils as water-borne nitrogen infiltrates into 

the soils and groundwater, and denitrification that occurs 

in highly-organic, saturated soils such as in wetlands 

where oxygen is low. 

Most BMPs are considered structural, and may 

be highly engineered “package” plants that can treat 

sewage or runoff, depending on scale and structure, 

or simple detention basins that allow sediments and 

adhered pollutants to settle out. “Artificial” wetlands are 

a good example of a more sophisticated BMP that takes 

advantage of natural processes, and may be created at the 

end of the stormwater pipe, or at edge of field. Structural 

BMPs are an important part of any strategy to limit 

reactive nitrogen loss to the environment. For example, 

The State of Iowa contains some of the most productive 

agricultural land in the world. Of the 36 million acres of 

land, 23 million acres are planted in corn or soybeans. 

Approximately 39 percent of the corn/soybean acres 

are drained with an estimated 800,000 miles of tile 

(Cutler, 2000). Each year, thousands of miles of new or 

replacement tile are installed. This drainage network is 

responsible for the conveyance of 90 percent (Crampton 

et al., 2006) of the nitrate that appears in Iowa’s surface 

waters. Control of nitrogen discharge from drainage 

tile will be needed to limit reactive nitrogen loss to the 

environment from agriculture.

Various approaches have been proposed with varying 

degrees of success. Constructed wetland, bio-reactors, 

and drainage tile encapsulated with biomass have 

proven to reduce 90 percent of the reactive nitrogen 

loss to the environment (Blowes et al., 1994). It would 

seem reasonable to require any new or replacement 

drainage tile to implement a control strategy at the time 

of construction and a retrofit program for the remaining 

drainage systems. With an average of two years to pay 

back the cost of drainage systems due to increased crop 

yield, the added cost for nitrogen control does not seem to 

be unreasonable.

Non-structural BMPs are often preservation actions, 

as discussed earlier, or activities that prevent pollutants 

from entering the waste stream such as street sweeping or 

fertilizer limitation.

Engineered and restored wetlands to decrease 

NO3

-

loading of aquatic systems

The construction and/or restoration of wetlands have 

received considerable attention in the past two decades 

as a conservation method. Such an approach has several 

positive attributes including promoting denitrification in 

watersheds containing or receiving Nr, flood protection, 

habitat preservation, and recreational potential (Hey and 

Philippi, 1995). In the upper Mississippi basin optimum 

siting of wetlands could result in as much as 0.4Tg of NO3

-

converted to N2 (Mitsch et al., 1999, 2001; Hey, 2002).

Much of the nitrate leached from agricultural fields 

could be removed from drainage water in natural, 

created, or restored wetlands. Nitrate removal from the 

water column in wetlands is performed by plant uptake, 

sequestration in the soils, and microbial transformation 

that includes immobilization and denitrification. Plant 

uptake and microbiological immobilization result in 

temporary storages in the system since most nitrogen 

will eventually return to the wetland via plant death and 

decomposition. In contrast, denitrification can constitute 

a real nitrogen sink because NO3

-

 is converted mainly 

to N2 that is emitted to the atmosphere (Clement et al., 

2002). As discussed in Section 4.7, the potential for the 

formation of N2O is of concern if such systems are not 

operated properly.

In addition to preserving existing wetlands, there are 

two other basic approaches that utilize wetlands to reduce 

the Nr and other nutrients reaching rivers, streams, and 

vulnerable downstream coastal systems. These approaches 

are: 1) creation and restoration of ecosystems, principally 

wetlands and riparian forests, between farms and adjacent 

ditches, streams and rivers; and 2) diversion of rivers into 

adjacent constructed and restored wetlands all along the 

river courses (Mitsch and Jorgensen, 2004; Mitsch and 

Gosselink, 2007).
70

The Committee notes that if wetlands can be 

economically and effectively restored where croplands 

now exist on hydric soils within the 100-year floodplain, 

this may be an important NO3

-

 control mechanism. 

Cropland on hydric soil in the floodplain occupy about 

6.9 million acres (2.8 million hectares). If this area and its 

wetlands were given back to the Mississippi River, over 

a million tons of NO3

-

 -N would be annually removed or 

prevented from reaching the Gulf of Mexico (Mitsch et 

al., 1999; Hey, 2002; Hey et al., 2004). To give scale to the 

solution needed, restoration of over 4.9 million acres (2 

million hectares) of wetlands is needed in the MississippiOhio-Missouri (MOM) river basin to reduce the nitrogen 

load to the Gulf of Mexico sufficiently to ensure a 

reduction in the size of the hypoxic zone in the Gulf of 

Mexico (Mitsch et al., 2001). 

At a series of workshops on restoration of the MOM 

river basin in 2003-2004 (Day et al., 2005; Mitsch and 

Day, 2006), scientists and managers were asked to 

focus on needed research and chokepoint opportunities 

for managing N in that basin. They concluded that a 

major, interdisciplinary research program (as a leadin to the actual restoration of wetlands and rivers) was 

needed, with sufficient funding, study sites, and time 

to reduce remaining uncertainties about the efficacy of 

wetlands to solve pollution problems related to N. It 

was recommended that, to implement this program, 20 

to 30 full-scale, existing and new agricultural/wetland 

demonstration projects should be located throughout the 

country and instrumented to study agricultural runoff into 

wetlands in a variety of soil conditions. Pilot and full-scale 

studies of diversions into riparian systems along river 

channels were recommended in order to determine the 

effectiveness of these systems for nutrient removal. The 

Committee notes that these research and demonstration 

projects have not been undertaken, and that there is a 

continuing need for this work.

Further illustration of the use of wetlands as a tool for 

Nr management tool is presented in Appendix C, Water 

Quality Trading in the Illinois River Basin (D. Hey, 

personal communication).

Technical controls (control points) on transfer 

and transformation of atmospheric emissions of 

Nr in and between environmental systems: NH3

Newly fixed Nr is produced biologically or added as 

fertilizer to meet the demand for food and fiber production. 

Much of the N is used in cereal crop production and cereal 

crops are then used to feed livestock. The new Nr is then 

recycled through the livestock production system where 

it becomes again susceptible to losses to the atmosphere 

as ammonia and NOx, and is available for additional N2O 

production and movement into aquatic systems as NH4

+

and NO3

-

. 

The bulk of the N fed to livestock ends up in manure, 

and where this manure (approximately one-half in urine 

and one-half in feces) is produced, there is often a much 

greater supply than can be efficiently or economically 

used as fertilizer on crops. For large concentrated animal 

feeding operations there is considerable expense associated 

with disposal of the manure. Various storage systems have 

been developed to deal with this excess manure, the most 

interesting of which, from the standpoint of integrated 

policy on N, convert the urea to N2. The fraction of 

manure N that can be and is converted to N2 remains a 

major unanswered scientific or technical question.

The NRC (2003) noted the paucity of credible data on 

the effects of mitigation technology on rates and fates of 

air emissions from CAFOs. The report did, however, call 

for the immediate implementation of existing atmospheric 

emission technology. The NRC (2003) also called for a 

mass balance approach in which the losses of N species 

such as NH3, NO, N2, and N2O are expressed as a fraction 

of the total N loss. Quoting from the NRC report:

Storage covers for slurry storage tanks, anaerobic 

lagoons, and earthen slurry pits are being studied 

as a method to decrease emissions from those 

containments. Both permeable and impermeable 

covers are being studied. Tested covers range from 

inexpensive material such as chopped straw (on slurry 

containments only) to more expensive materials such 

as high density polyethylene. Covers can decrease 

emissions from storage but their net effect on 

emissions from the system is conditional on how the 

effluent is used on the farm.

Anaerobic digestion in closed containment has been 

studied for many types of applications. Anaerobic 

digestion is the process that occurs in an anaerobic 

lagoon. When conducted in closed vessels, gaseous 

emissions including methane, carbon dioxide and 

small amounts of other gasses (possibly ammonia, 

hydrogen sulfide, and VOCs) are captured and can be 

burned for electricity generation, water heating, or 

simply flared. The in-ground digester being tested on 

a swine farm in North Carolina is an example of the 

ambient temperature version of this technology (there 

are also mesophilic and thermophilic designs). The 

concentration of ammonia remaining in effluent from 

that digester is higher than the concentration in lagoon 

effluent and can be volatilized once exposed to air. 

Recent research (e.g., Bicudo et al., 2004; Funk et al., 

2004a,b; Shores et al., 2005) demonstrates reduction in 

NH3 emissions after a permeable cover was installed. 

Miner et al. (2003) reported that a polyethylene cover 

can reduce NH3 emissions by ~80%, but it is not clear 

what fraction of that N was converted to N2. Harper et 

al. (2000) reported that in a well-managed swine lagoon 

denitrification N2 losses can be equivalent to N lost as 

NH3, in other words about 50% efficiency. Kermarrec 

et al. (1998) reported that sawdust litter helped reduce 

NH3 emissions from pig manure with 44-74% of manure 
71

N converted to N2, but greater than 10% of the manure 

N was released as N2O. Sommer (1997) found that NH3

was emitted from cattle and pig slurry tanks at the rate 

of 3.3 kg N m2yr-1 until covered with straw. After straw 

application NH3 emissions were below detection limit. 

Mahimairaja et al. (1994) reported that NH3 volatilization 

was reduced by 90-95% under anaerobic conditions. 

Section 2.2.4 contains a discussion of best management 

practices to minimize NH3 emissions from livestock 

waste, and presents finding and recommendation 6 on the 

need for a framework for manure management.

Technical controls (control points) on transfer 

and transformations of atmospheric emissions of 

Nr in and between environmental systems: NOx

As previously discussed, a major contributor to Nr 

in the atmosphere is fossil fuel combustion. During the 

combustion process NOx (NOx = NO + NO2) are released 

to the atmosphere. Globally the production of NOx has 

accelerated in the last few decades, primarily through the 

increase in fossil fuel combustion (Galloway et al., 1995; 

2008). With this increase in emissions from ~5 Tg N in 

1940 to ~ 25 Tg N in 2005, combustion of fossil fuels 

accounted for about 50% of the total global NOx emissions 

for 1990. Of the anthropogenic sources, fossil fuel, 

aircraft, biomass burning, and part of the soil emission are 

most important (Holland et al., 1997). Although global 

NOx emissions continue to increase, these emissions are 

declining in the U.S. (see Section 2.2.1).

Nitrogen oxide is formed during combustion by three 

mechanisms:

n Thermal NOx where N2 and O2 combine at high 

temperatures (thermal pathway dominates at 

temperatures greater than approximately 1500 C) to 

form NO through the Zeldovich mechanism

n Fuel NOx where nitrogen from a fuel (e.g., coal and 

biofuels) is released as some intermediate and then 

combines with O2 to form NO

n Prompt NOx where N2 reacts with hydrocarbon radicals 

in flames, forming various compounds including 

hydrogen cyanide and other cyano radicals. These 

in turn form NOx. Contributions of prompt NOx are 

usually low as compared to fuel NOx.

There are several ways to control emissions of NOx. 

The most common controls are on coal-fired electric utility 

generators and those are discussed below. Following the 

discussion of electric utility generator controls, or external 

combustion systems, there is a discussion on internal 

combustion controls.

Reduction of the temperature limits the kinetics of 

the N/O2 reaction. Temperature can be controlled by 

using a fuel-rich mixture versus fuel lean. In this case 

the reactions take place at lower temperatures. Fuelrich mixtures also reduce the amount of O2 available 

for reaction and there are changes in the chemical 

mechanisms which limit the oxidation of N2. If fuellean mixtures are used for temperature control, while the 

temperature is lower there is a significant amount of O2

present. Typically, in external combustion systems controls 

are implemented by using less excess air and using staged 

combustion. In addition, flue-gas recirculation (FGR) is 

used to lower the temperature. Low-NOx burners operate 

under the principle of internally staging the combustion. 

To reduce fuel NOx, air and fuel staging are used to reduce 

the peak temperature where air and fuel are admitted in 

separate locations. 

Chemical reduction of NOx is also possible. These 

methods include: selective non-catalytic reduction 

(SNCR), selective catalytic reduction (SCR), and fuel 

reburning. SNCR is an add-on technology where urea or 

NH3 is injected in a controlled temperature zone to allow 

the reduction of NOx. SCR is also an add-on technology 

where the flue gas must pass through a catalyst bed to 

allow reaction between ammonia and NOx. Care must 

be taken with both technologies to avoid NH3 slip. Fuel 

reburning requires the injection of a fuel to create a zone 

where NOx is reduced to N2. Low NOx burners may also 

use an internal fuel reburning to reduce the NOx. 

For internal combustion engines, the same mechanisms 

discussed above are used in a variety of different 

ways, since these systems are using high pressure and 

predominately have thermal NOx versus fuel NOx

formation. Most technologies involve the need to reduce 

the peak temperature and duration of high temperatures 

of the combustion zone. For example, gas turbines utilize 

low NOx burners, while spark ignition engines utilize a 

three-way catalyst which requires less than 0.5% O2. In 

this case, additional NOx is reduced by utilizing unburned 

fuel as a reagent over the catalyst for chemical reduction 

of NOx. It should be noted, however, that a side reaction 

for the three-way catalyst system produces ammonia. NOx

emissions can be reduced in diesel engines by delaying the 

injection of the fuel, and by retarding the timing in sparkignited engines. Engines also use exhaust gas recirculation 

(EGR) to reduce the peak temperatures. Recent road side 

studies have indicated high efficiency (~90%) for NOx

removal from the American light-duty fleet (Bishop and 

Stedman, 2008). 
72
73

6

SAB Findings and Recommendations for 

Nr Data Collection, Risk Management, 

and Research

This concluding chapter contains the Integrated 

Nitrogen Committee’s findings and its recommendations to 

EPA. Section 6.1 discusses the need for a comprehensive 

program to monitor reactive nitrogen. Section 6.2 provides 

the Committee’s overarching recommendations to EPA. 

Section 6.3 contains suggestions for near-term actions 

that might be taken by EPA and other management 

agencies to decrease Nr entering the environment from 

various sources. Section 6.4 contains specific findings 

and recommendations corresponding to each of the 

Committee’s four study objectives.

The first objective of the study was to identify and 

analyze, from a scientific perspective, the problems Nr 

presents in the environment and the links among them. 

To accomplish this objective, the Committee examined 

the flows of Nr within the food, fiber, feed and bioenergy 

production systems and developed lands in the U.S., 

paying special attention to the locations within each 

of these systems where Nr is lost to the environment. 

The same process was employed for fossil fuel energy 

production but, since all the Nr formed and released 

during energy production is lost to the environment, the 

Committee identified the important energy producing 

sectors that contribute to Nr emissions. The Committee 

found that agriculture and domestic use of fertilizers 

to produce food, feed, and fiber (including bioenergy 

and BNF) and combustion of fossil fuels are the largest 

sources of Nr released into the environment in the U.S. 

The Committee also examined the fate of the Nr 

lost to the environment, estimated the amount stored in 

different systems (e.g., forest soils) and tracked Nr as 

it is transferred from one environmental system (e.g., 

the atmosphere) to another (e.g., terrestrial and aquatic 

ecosystems). Source and fate analyses set the stage for 

identifying the environmental and human health problems 

Nr presents, and the links among them. Using the nitrogen 

cascade, the Committee identified the impacts Nr has 

on people and ecosystem functions as it moves through 

each system and contributes to adverse public health and 

environmental effects, including photochemical smog, 

nitrogen-containing trace gases and aerosols, decreased 

atmospheric visibility, acidification of terrestrial and 

aquatic ecosystems, eutrophication of coastal waters (i.e., 

harmful algal blooms, hypoxia), drinking water concerns, 

freshwater Nr imbalances, GHG emissions and subsequent 

climate change, and stratospheric ozone depletion. 

The second objective of the study was to evaluate 

the contribution an integrated N management strategy 

could make to environmental protection. To accomplish 

this objective the Committee identified actions that 

could be taken to better manage Nr. These actions take 

into account the contributions of all Nr sources and 

chemical species that adversely impact human health 

and environmental systems, and the need to ensure that 

solving one problem related to Nr does not exacerbate 

another problem or diminish ecosystem services that 

support societal demands. 

The third objective of the study was to identify 

additional risk management options for EPA’s 

consideration.In addressing this objective, the Committee 

identified four major goals for management actions that 

collectively have the potential to decrease Nr losses to 

the environment by about 25%. Decreasing Nr emissions 

by these actions will result in further decreases in 

Nr-related impacts throughout the nitrogen cascade. 

The Committee has suggested a number of ways to 

attain these management goals, including conservation 

measures, additional regulatory steps, voluntary actions, 

application of modern technologies, and end-of-pipe 

approaches. The Committee notes that these are initial but 

significant actions; however, others should be taken once 

the recommended actions are completed and assessed, 

and further opportunities are explored in an adaptive 

management approach. 

The fourth objective of the study was to make 

recommendations to EPA concerning improvements in 

Nr research to support risk reduction. The Committee has 

provided numerous recommendations for additional Nr 

research to support risk reduction activities. 

6.1. Need for Comprehensive Monitoring 

of Nr

In previous sections of this report the Committee 

has discussed the importance of monitoring reactive 

nitrogen in the environment. The Committee recommends 

establishing a program for comprehensive monitoring 

of the multiple forms of reactive nitrogen as both 

stocks and flows as they pass through different media 

and ecosystems. There are two major reasons for this 

overarching recommendation. The first purpose of 

monitoring is to provide the observational data on trends 

that will inform research into the complexity of the 
74

nitrogen cascade to better identify the most effective 

intervention points to reduce damage to human health and 

the environment by reactive nitrogen. The second need 

for monitoring is to be able to assess the effectiveness 

of policy interventions over time, and to apply the 

principles of adaptive management. As it becomes clear 

which strategies and policy instruments are effectively 

reducing the amounts of Nr entering the environment, 

and which are ineffective, it will be necessary to modify 

those interventions in response to the monitoring data. As 

conditions change (e.g., shifts in the technology of electric 

power production, new fuels for transportation, changing 

land use patterns and climate change), the nitrogen 

cascade will be modified, and the relative importance 

of sector specific policies will change. Only through 

comprehensive monitoring will it be possible to manage 

Nr effectively. 

Finding 20: The Committee has determined that an 

integrated approach to monitoring that includes multiple 

media (air, land, and water) components and considers 

a suite of environmental and human concerns related to 

reactive nitrogen in the environment (e.g., Nr effects, 

climate change, human health) is needed. Some of the 

phenomena presented in this report need more definition 

and verification but, more importantly, as controls are 

brought to bear on Nr, improvements need to be measured 

to verify and validate successful management strategies. 

If the desired improvements are not realized as shown by 

the collected data, corrective measures will be required. 

Such an adaptive approach acknowledges the likelihood 

that management programs will be altered as scientific 

and management understanding improve. 

Recommendation 20: The Committee recommends 

that EPA initiate discussions and take action to develop a 

national, multi-media monitoring program that monitors 

sources, transport and transition, effects using indicators 

where possible, and sinks of Nr in keeping with the 

nitrogen cascade concept. This comprehensive program 

should build upon existing EPA and state initiatives as 

well as monitoring networks already underway in other 

federal agencies such as the U.S. Geological Survey 

programs and the NADP effort.

6.2. Overarching Recommendations

Human activities have significantly increased the 

introduction of Nr into the U.S. environment and, through 

radical alterations of land use, have eliminated many 

of the natural features that once may have provided 

pollutant treatment. While there have been significant 

benefits resulting from food production, there have also 

been, and continue to be, major risks to the health of both 

ecosystems and people due to the introduction of Nr into 

the nitrogen cascade. 

In its 1990 report, Reducing Risk, the Science Advisory 

Board recommended that the EPA increase its efforts 

to integrate environmental considerations into broader 

aspects of public policy in as fundamental a manner as 

are economic concerns. Other federal agencies often 

affect the quality of the environment, e.g., through 

the implementation of tax, energy, agricultural, and 

international policy, and EPA should work to ensure 

that environmental considerations are integrated, where 

appropriate, into the policy deliberations of such agencies. 

In the current era of increasing responsibilities without 

commensurate budgets, intergovernmental cooperation, 

partnerships and voluntary programs have become vital 

tools for agencies needing to stretch their resources to 

fulfill their missions. 

Optimizing the benefits of Nr, and minimizing its 

impacts, will require an integrated nitrogen management 

strategy that not only involves EPA, but also 

coordination with other federal agencies, the States, the 

private sector, universities, and a strong public outreach 

program. The Committee understands that there are real 

economic costs to the recommendations contained in this 

report. For each recommendation there will of necessity 

be tradeoffs derived from the varying cost-effectiveness 

of different strategies. 

The Committee makes four overarching 

recommendations:

Overarching Recommendation 1

The Committee recommends an integrated approach 

to the management of Nr. This approach will likely use 

a combination of implementation mechanisms. Each 

mechanism must be appropriate to the nature of the 

problem at hand, be supported by critical research on 

decreasing the risks of excess Nr, and reflect an integrated 

policy that recognizes the complexities and tradeoffs 

associated with the nitrogen cascade. Management efforts 

at one point in the cascade may be more efficient and cost 

effective than control or intervention at another point. 

This is why understanding the nature and dynamics of the 

N cascade is critically important.

Overarching Recommendation 2

The framing of the reactive nitrogen cascade provides 

a means for tracking nitrogen as it changes form and 

passes through multiple ecosystems and media. This 

complexity requires the use of innovative management 

systems and regulatory structures to address the 

environmental and human health implications of the 

massive amounts of damaging forms of Nr. It is difficult 

to create fully effective regulations de novo for such a 

complex system so we recommend utilizing adaptive 

management to continuously improve the effectiveness 

and lower the cost of implementation policies. This in 

turn will require a monitoring system that will provide 

feedback on the effectiveness of specific actions taken to 

lower fluxes and concentrations of Nr. 

Overarching Recommendation 3

EPA should form an intra-Agency Nr management 

task force that will build on existing Nr research and 
75

management capabilities within the Agency. This 

intra-Agency task force should be aimed at increasing 

scientific understanding of: (1) Nr impacts on terrestrial 

and aquatic ecosystems, human health, and climate,(2) 

Nr-relevant monitoring requirements, and (3) the most 

efficient and cost-effective means by which to decrease 

various adverse impacts of Nr loads as they cascade 

through the environment. 

Overarching Recommendation 4

Successful Nr management will require changes in 

the way EPA interacts with other agencies. To coordinate 

federal programs that address Nr concerns and help 

ensure clear responsibilities for monitoring, modeling, 

researching and managing Nr in the environment, the 

Committee recommends that EPA convene an Interagency Nr management task force. It is recommended 

that the members of this inter-agency task force include at 

least the following federal agencies: U.S. Department of 

Agriculture, U.S. Department of Energy, U.S. Department 

of Housing and Urban Development, U.S. Department 

of Transportation, National Oceanic and Atmospheric 

Administration, U.S. Geological Survey, U.S. Forest 

Service, and Federal Emergency Management Agency . 

This task force should coordinate federal programs that 

address Nr concerns and help ensure clear responsibilities 

for monitoring, modeling, researching, and managing Nr 

in the environment. The EPA Office of International and 

Tribal Affairs should work closely with the Department of 

State to ensure that EPA is aware of international efforts 

to control Nr and is developing national strategies that 

are compatible with international initiatives.

Similar recommendations for coordination and joint 

action among and between agencies at both state and 

federal levels have been made in the National Research 

Council’s recent reports on the Mississippi Basin 

(NRC, 2008b, 2009). These intra and inter-agency Nrmanagement task forces should take a systems approach 

to research, monitoring, and evaluation to inform public 

policy related to Nr management. The Committee 

proposes that the intra and inter-agency task forces 

coordinate the following activities to implement a systems 

approach to Nr management.

Development of methods 

Implementation of a systems approach will require 

development of methods to facilitate various aspects of Nr 

management. These include methods for: (1) establishing 

and evaluating proposed Nr budgets; (2) using life cycle 

accounting approaches for Nr management; (3) gathering 

and using data on N fertilizer use and other Nr sources 

and fluxes as the basis for informed policies, regulations 

and incentive frameworks for addressing excess Nr loads; 

(4) evaluating the critical loads approach to air and water 

quality management; (5) identifying and using indicators 

of excess Nr’s economic damage and effects on human 

health and the environment; and (6) using systems-based 

approaches for controlling Nr releases to the environment.

Implementing best management practices 

(BMPs)

It will be necessary to improve the scientific 

understanding of BMPs that can be used for specific 

applications to manage Nr. In particular, this includes 

better scientific understanding of: (1) Nr requirements 

in agriculture to ensure adequate food, feed, fiber, and 

bioenergy feedstock supply while also avoiding negative 

impacts on the environment and human health; (2) Nr 

requirements for urban landscapes (e.g., residential 

and commercial) and their maintenance while avoiding 

negative impacts on the environment and human health; 

(3) planning and pollution prevention, including low 

impact development and natural ecosystem service 

preservation; (4) use of natural land features and 

attributes, such as wetland preservation and enhancement, 

natural soil profiles, and buffer strips; and (5) improved 

removal of Nr from sewage waste steams at both largescale wastewater treatment facilities and individual 

subsurface (septic) systems. In addition, proactive 

extension and technology transfer approaches will need to 

be established to facilitate adoption of BMPs.

Developing appropriate tools and metrics for 

assessing impact from adoption of best management practices

Assessment activities will also be an important element 

of the systems approach to managing Nr. These activities 

should include: (1) quantifying the effectiveness and 

impact of policies and regulations focused on reduction 

of negative environmental impacts from Nr; (2) assessing 

combined carbon (C) and Nr effects on terrestrial and 

aquatic ecosystems; (3) assessing indicators/endpoints, 

costs, benefits, and risks associated with impairment of 

human health and decline and restoration of ecosystem 

services; (4) reviewing existing and proposed legislation 

for purposes of better integrating or designing regulatory 

activities that recognize the nitrogen cascade or 

streamlining procedures for enacting Nr risk reduction 

strategies; and (5) evaluating economic incentives, 

particularly those that integrate air, aquatic, and land 

sources of excess Nr.

Education, outreach, and communication

It will be necessary to develop new education, 

outreach, and communication initiatives. As discussed in 

this report, this includes a range of targeted outreach and 

education programs to manage Nr and achieve desired 

environmental outcomes. 

6.3. Near-term Management Goals

The Committee puts forward four goals for actions 

that could be taken by EPA or other management agencies 

to decrease Nr entering the environment from various 

sources. We believe these goals can be attained over the 

near term (approximately 10-20 years) using existing and 

emerging technologies and practices. These suggestions, if 

implemented, have the potential to reduce total Nr loadings 
76

to the environment in the U.S. by approximately 25% 

below current levels. The Committee believes that these 

represent realistic and attainable near-term goals, however 

further reductions are undoubtedly needed for many 

N-sensitive ecosystems and to ensure that health-related 

standards are maintained. The Committee understands that 

actual policy decisions on the implementation of programs 

to limit Nr releases to the environment may differ from 

those listed below for a variety of reasons, but believes that 

an aggressive level of action, as represented by these goals, 

is critical given the growing demand for food and fiber 

production and energy use from population pressures and 

economic growth. The rationale for these goals is set forth 

below, along with recommended management options for 

achieving the goals.

Management Goal 1. Controls on NOx emissions 

from mobile and stationary sources

The Clean Air Act (1970) and its Amendments (1990) 

have resulted in NOx emissions that are less than 50 

percent of what they would have been without existing 

controls. While this is an admirable accomplishment, 

there is still a need to seek improvements. NOx emissions 

are an order of magnitude greater than at the beginning 

of the twentieth century. As a consequence, there 

remain significant negative impacts on both humans 

and ecosystems. In 2002, coal-fired utilities generated 

approximately 1.3 Tg N annually (see Figure 3). If all 

coal-fired plants used state-of-the-art NOx controls, this 

number could be reduced by 0.6 Tg N/yr (calculations 

performed by Cohen, 2008); in fact, 2008 emissions have 

been reduced by 0.3 Tg N/yr from 2002 levels (see Figure 

3), so in essence, half the reduction has already been 

accomplished. The EPA should continue to reduce NOx

emissions from major point sources, including electric 

generating stations and industrial sources, expanding 

the use of market mechanisms such as cap and trade. 

Under this scenario, it is likely that high-efficiency, lowemission power plants will be built for energy needs. 

Some controls on NOx emissions are implemented only 

in the ozone season (May to September) (U.S. EPA 

2009c). To protect welfare and avoid adverse effects 

on ecosystems, NOx emissions controls should be 

implemented year round.

For mobile sources, emissions for highway and 

off-highway sources are approximately 2.2 Tg N/yr 

and 1.2 Tg N/yr, respectively. EPA is in the process of 

implementing a number of regulations that will reduce 

NOx from mobile sources (see Appendix F). The 

implications of these recent regulations are not reflected 

in the quantitative analysis presented in this report. EPA 

projects future year decreases in emissions (see Figure 

5 in Section 2.2.1). However, better controls are needed 

for on-road heavy duty diesel vehicles and off-road 

vehicles, which include locomotives, construction, farm, 

landscaping equipment, and marine vehicles. For these 

off-road vehicles, 80-90% NOx removal is technically 

achievable (deNevers, 1995; Koebel et al., 2004). 

Assuming a 40% reduction for these sources, there is a 

potential reduction of 1.4 Tg. The total reduction for both 

mobile and stationary sources is then approximately 2 

Tg N/yr. Part of achieving such levels of compliance will 

require the implementation of inspection and maintenance 

programs or road-side monitoring. 

The Committee cautions, however, that achieving 

such a goal may be inadequate for many areas to meet 

the new 60 to 70 ppb ozone standard recommended 

by the EPA Clean Air Scientific Advisory Committee 

(CASAC) (U.S. EPA CASAC, 2008) or even the 75 ppb 

standard currently promulgated. Additional measures 

such as increasing the role of solar- and wind-generated 

electricity, wider use of hybrid and electric cars, and 

public transit conducive to energy conservation and 

reduced emissions should be promoted.

Management Goal 1. The Committee estimates 

that if EPA were to expand its NOx control efforts 

for emissions of mobile sources and power plants 

and include implementation of year round controls 

on stationary sources to protect welfare and avoid 

adverse effects on ecosystems, a 2.0 Tg N/yr decrease 

in the generation of reactive nitrogen could be 

achieved. It is believed that coal-fired utilities could 

experience a reduction of 0.6 Tg N/yr. Since 2002, 

emissions have already been reduced by at least 0.3 

Tg N/yr; hence, this represents an additional 0.3 Tg 

N/yr. Approximately 3.4 Tg N/yr can be attributed to 

mobile sources (highway, off-highway). Assuming a 

conservative 40% reduction, an additional 1.4 Tg N/yr 

could be reduced.

Management Goal 2. Nr discharges and emissions from agricultural lands and landscapes

Section 5.3.4 of this report reviews the various 

methods that can be used to improve Nr management in 

agricultural systems. The Committee estimates that crop 

N-uptake efficiencies can be increased by up to 25% over 

current levels through a combination of knowledge-based 

practices and advances in fertilizer technology (such as 

controlled release and inhibition of nitrification). Crop 

output can be increased while reducing total Nr by up 

to 20% of applied synthetic fertilizers, approximately 

2.4 Tg N/yr below current levels of Nr additions to the 

environment. These are appropriate management goals 

with today’s available technologies. Further progress is 

possible through expanded research programs. 

The Committee is concerned about current policies 

and practices governing biofuel development. Acreage 

devoted to corn production has increased substantially 

for corn based ethanol production during the past several 

years (with nearly one-third of the crop currently devoted 

to bioethanol production), with fertilizer nitrogen use on 

corn increasing by at least 10% (an additional 0.5 Tg N/

yr), largely to meet biofuel feedstock crop demand. In the 
77

absence of Nr controls and a failure to implement best 

practices, current biofuels policies will make it extremely 

difficult to reduce Nr transfers to soils, water and air 

(Simpson et al., 2008). Integrated management strategies 

will be required. 

The Committee also notes with concern the increase 

of N2O in the atmosphere. The Committee believes that 

GHG emissions trading will provide both opportunities 

and challenges for mitigating Nr environmental and health 

impacts. Policies and regulations should consider how to 

reward reductions of N-related GHG. Biofuel subsidies that 

accurately account for Nr contributions to GHG emissions, 

certification of individual biofuel plants for GHG impact, 

and rewards for farmers who reduce N2O emissions 

are examples of how an integrated strategy can reduce 

agricultural GHG impacts. For additional production of 

liquid biofuels beyond the grandfathered amount in the 

Energy Independence and Security Act (EISA), EPA has 

the power to exercise some controls on N2O emissions 

through the life cycle GHG accounting requirements. It is 

the opinion of the Committee that Section 204 of the EISA 

calling on the Agency to adopt a life cycle approach to the 

assessment of future renewable fuel standards is a positive 

step toward a comprehensive analysis. 

Section 5.3.4 of this report reviews methods of 

controlling Nr from landscape runoff through the use of 

natural or restored wetlands, urban areas, and through 

the use of best management practices. The Committee 

finds that flows of Nr into streams, rivers, and coastal 

systems can be reduced by approximately 20% (~1 Tg N/

yr) through improved methods of landscape management 

and without undue disruption to human commercial and 

aesthetic activities. 

Management Goal 2. The Committee estimates that 

crop N-uptake efficiencies can be increased by up to 

25% over current practices through a combination of 

knowledge-based practices and advances in fertilizer 

technology amounting to ~2.4 Tg N/yr below current 

amounts of Nr additions to the environment. The 

Committee further estimates that excess flows of 

Nr into streams, rivers, and coastal systems can be 

decreased by approximately 20% (~1 Tg N/yr) through 

improved landscape management and without undue 

disruption to agricultural production. 

Management Goal 3. Ammonia emissions from 

livestock management and manure handling

In spite of gains made over the last several decades in 

decreasing the amount of NOx emitted from stationary 

and mobile combustion sources, the total amount of Nr 

released into the atmosphere has remained relatively 

constant. This is related to the essentially unregulated 

release of ammonia from livestock operations. 

As discussed in Section 2.2.3, at the present time, 

fewer livestock are required to produce more animal 

products than in the past. For example, since 1975 milk 

production has increased linearly at the rate of ~ 180 

kg milk per cow /yr while milk cow herd population 

decreased at the rate of ~69,000 head per yr (i.e., the 

60% greater amount of milk produced in 2006 compared 

to 1970 required 25% fewer cows). Animal inventories 

declined by 10% for beef brood cows from 36 million 

head in 1970 to 33 million head in 2006, and the 

inventory of breeder pigs and market hogs declined 

8% from 673 million head to 625 million head in the 

same period, even with similar or greater annual meat 

production. These trends resulted from greater growth 

rates of animals producing more meat in a shorter 

amount of time. In 1970, broilers were slaughtered after 

80 days on feed at 1.7 kg live weight, but by 2006 the 

average weight was 2.5 kg after only 44 days on feed. 

These trends in requiring fewer animals to produce 

more animal food products through improved diet and 

increased production efficiency will continue. 

Implementation of improved methods of livestock 

management and manure handling and treatment to 

decrease NH3 emissions that have been developed 

since 1990 will further decrease ammonia and other 

gases and odor emissions. For example, sawdust litter 

helps decrease NH3 emissions from pig manure with 

44-74% of manure N converted to N2. Storage covers 

for slurry storage tanks, anaerobic lagoons, and earthen 

slurry pits decrease emissions from those containments. 

Anaerobic digestion in closed containment has been 

studied for many types of applications. Recent research 

demonstrates reduction in NH3 emissions after a 

permeable cover was installed (e.g., a polyethylene cover 

decreased NH3 emissions by ~80%). A well managed 

swine lagoon can denitrify approximately 50% of the 

excreted N to N2. Recently engineered developments 

utilizing closed loop systems (Aneja et al., 2008a) 

substantially reduce atmospheric emissions of ammonia 

(> 95%) and odor at hog facilities. Based upon recently 

demonstrated reduction of NH3 emissions from swine 

and poultry production, a moderate reduction of 50% 

from 1990 NH3 emission estimates for swine and poultry 

production should be attainable (Table 17). Because 

of the larger land area involved in dairy and beef 

production and the lack of effort that has been exerted in 

mitigating NH3 emissions, a more modest and reachable 

goal of decreasing NH3 emissions by 10% through 

improvements in animal diet and manure management is 

proposed (Table 17).

Management Goal 3. The Committee estimates that 

livestock-derived NH3 emissions can be decreased by 

30% (a decrease of 0.5 Tg N/yr) by a combination of 

BMPs and engineered solutions. This is expected to 

decrease PM2.5 by ~0.3 μg/m3 (2.5%), and improve 

health of ecosystems by achieving progress towards 

critical load recommendations. Additionally we 

estimate that NH3 emissions derived from fertilizer 

applications can be decreased by 20% (a decrease 
78

of ~0.2 Tg N/yr), through BMPs that focus on 

improvements related to application rate, timing, 

and placement.

Management Goal 4. Discharge of Nr from 

developed lands and point sources 

National loadings of Nr to the environment from 

public and private wastewater point sources are relatively 

modest in comparison with global Nr releases; however, 

they can be important local sources with associated 

impacts, especially in highly-populated watersheds. The 

Committee has estimated that sewage containing Nr from 

human waste contributes 1.3 Tg N/yr to the terrestrial 

inputs of nitrogen (Table 1). 

The Committee has also estimated that turf fertilizer 

usage contributes 1.1 Tg N/yr to terrestrial inputs, a load 

that could potentially be cut by about one third (Section 

2.2.4). The Committee did not provide estimates for 

general stormwater and nonpoint source runoff nitrogen 

load reductions specific to developed or urban areas 

– runoff concentrations and loads are highly variable 

reflecting geographic and climatic conditions throughout 

the U.S. and equally variable removal efficiencies from 

standard treatment BMPs. This is shown in a summary 

of the International Stormwater Best Management 

Practices Database (Geosyntec Consultants, Wright 

Water Engineers, Inc., 2008). However, most BMPs are 

effective because they provide the beneficial biochemical 

conditions of wetlands, the biophysical controls described 

in Section 5.3.4 and Appendix C. These benefits, and the 

application of BMPs, are recommended in overarching 

recommendation 4, as well as in the preceding 

Management Goal 2 as applied to agricultural lands. 

Similar stormwater and nonpoint source management 

benefits specific to developed lands should be anticipated 

with BMP application in those areas.

Denitrification processes as applied to human 

waste at sewage treatment plants are well-studied and 

growing in application. Performance of these engineered 

solutions, collectively referenced as biological nitrogen 

removal (BNR), can be more rigorously controlled 

than for stormwater and nonpoint source BMPs. Recent 

publications by the U.S EPA (2007f, 2008e,f ) have 

summarized the state of and the capability for nitrogen 

removal, and have reported that technologies to achieve 

effluent concentrations of 3 mg total nitrogen per liter 

(TN/L) or less are readily available. However, plant 

capacity and design, wastewater characteristics, and 

climate conditions can all affect the ability of a facility 

to remove nitrogen. EPA’s review of 2003-2006 data 

for 16 facilities that remove nitrogen to varying degrees 

found a range of final effluent TN concentrations of 1.0 

to 9.7 mg/L, with an average of 5.6 mg/L. In general, 

very small facilities (<0.1 MGD) do not perform as 

well, with a final TN concentration ranging from 6-12 

mg/L. Treatment performance varied and exceeded 5 mg 

TN/L at some of the facilities. Given these conditions, 

and performance uncertainties, it seems reasonable to 

conclude that removal efficiencies in the range of 40-

60% below standard effluent nitrogen loads could be 

readily attained. Based on the human waste load of 1.3 

Tg N/yr, this would yield a decrease in total nitrogen 

load of between 0.5 and 0.8 Tg/yr. Using data provided 

by Maryland Department of the Environment (2006) and 

the Connecticut Department of Environmental Protection 

(2007), two states that have promoted nitrogen removal 

technologies as solutions to coastal eutrophication, EPA 

(2007f) has constructed cost estimates of upgrading the 

performance of sewage treatment plants ranging from 

$990,000 to $1.74 million per MGD treated. 

There are two funding sources of significance 

authorized in the CWA that are used to fund projects 

Estimate is based on livestock emissions of 1.6 Tg from Table 1.

Table 17: Estimates for potential decreases in NH3 emissions from livestock manure in the 

United States 

NH3 Source % of Total NH3 Tg NH3-N/yr emitted

Estimated Decrease of NH3

% Tg N/yr

Dairy 23.1 0.37 10 0.040

Beef 27.1 0.44 10 0.040

Poultry 27.5 0.44 50 0.220

Swine 17.5 0.28 50 0.140

Goat/sheep 1.6 0.03 10 0.003

Horse 2.9 0.05 10 0.005

Total 100 1.61 0.45
79

relevant to the control of Nr. Section 319 establishes 

state nonpoint source management programs to plan for 

and implement management measures that abate sources 

of nonpoint pollution from eight source categories, 

including both urban and agricultural sources; however, 

the CWA disallows use of 319 funds for NPDES 

permit requirements, so urban areas with stormwater 

permits do not qualify for Section 319 funding. Over 

the years, section 319 has made available, through 60% 

matching funds, over $1.6 billion in assistance. The 

much larger source of funding comes under Title VI of 

the CWA, which has provided over $24 billion (federal) 

for the construction of treatment facilities for point 

sources of wastewater over the past 20 years, although 

only a fraction of this amount has been dedicated to 

denitrification processes. Title VI “state revolving” loan 

funds can be used for stormwater management, as well as 

other water pollution management activities, but not all 

states have chosen to use funds beyond traditional sewage 

treatment plant infrastructure needs because of the large 

backlog of demand for those purposes.

In 2009, under the American Reinvestment and 

Recovery Act (ARRA), the CWA Clean Water State 

Revolving Fund (CWSRF) received a $4 billion boost 

for clean water infrastructure and the CWSRF for fiscal 

year 2010 was tripled over the prior years to two billion 

dollars. These additional funds not only provide for jobs 

creation, as intended by Congress, but provide states with 

resources to reduce the backlog of clean water projects, 

which also often include nutrient management needs. 

The ARRA funds also emphasized the use of CWSRF 

dollars for stormwater and nonpoint source management 

and energy savings under a “green infrastructure” 

requirement. A 20% set-aside for green infrastructure 

was a requirement of AARA CWSRF funding and was 

used widely for projects that included reductions in GHG 

emissions, land-based low impact development BMPs 

to reduce runoff and improve runoff quality, and other 

innovative practices to treat wastewater and runoff. A 

green infrastructure requirement is being continued in the 

fiscal year 2010 CWSRF allocation.

Management Goal 4. The Committee recommends 

that a high priority be assigned to increasing funding 

for nutrient management. We estimate that a decrease 

in Nr emissions from human sewage of between 0.5 

and 0.8 Tg N/yr can be achieved, with additional 

decreases likely with increased stormwater and 

nonpoint source BMP application support.

6.4. Summary of Specific Findings and 

Recommendations Corresponding to the 

Four Study Objectives

In this report the Committee has provided specific 

findings and recommendations to assist EPA in its 

understanding and management of nitrogen-related air, 

water, and soil pollution issues. The specific findings and 

recommendations corresponding to each of the four study 

objectives are summarized below. 

Study Objective #1: Identify and analyze, from a 

scientific perspective, the problems Nr presents 

in the environment and the links among them.

In general, the Committee finds that uncertainty 

associated with rapid expansion of biofuels, losses of Nr 

from grasslands, forests, and urban areas, and the rate and 

extent of denitrification have created the need to measure, 

model, and report all forms of Nr consistently and 

accurately. Addressing this need will decrease uncertainty 

in the understanding of the fate of Nr that is introduced 

into the environment and lead to a better understanding 

of the impacts of excess Nr on the health of people and 

ecosystems. This should be accomplished through a 

coordinated effort among cognizant federal and state 

agencies and universities. 

The Committee recommends that EPA routinely 

and consistently account for the presence of Nr in the 

environment in forms appropriate to the medium in which 

they occur (air, land, and water) and that accounting 

documents be produced and published periodically. 

Specific Findings:

n Rapid expansion of corn-ethanol production has the 

potential to increase N fertilizer use through expanding 

corn production and its associated N fertilizer inputs. 

Development of cellulosic ethanol industry will require 

cultivation for cellulosic crops, which will also require 

N fertilizer. Distillers grains are changing animal 

diets and affecting N recycling in livestock. Both 

have important consequences for the effective future 

management of Nr. (Finding #4 – also pertains to study 

objectives 2 and 4)

n Although total N budgets within all terrestrial systems 

are highly uncertain, Nr losses from grasslands and 

forests (vegetated) and urban (populated) portions 

of the N cascade appear to be higher, on a percent of 

input basis, than from agricultural lands. The relative 

amount of these losses ascribed to leaching, runoff, 

and denitrification are as uncertain as the N budgets 

themselves. (Finding #9)

n Denitrification of Nr in terrestrial and aquatic systems 

is one of the most uncertain parts of the nitrogen cycle. 

Denitrification is generally considered to be a dominant 

N loss pathway in both terrestrial and aquatic systems, 

but it is poorly quantified. (Finding #10 – also pertains 

to study objective 4)

n The Committee finds that there is a need to measure, 

compute, and report the total amount of Nr present in 

impacted systems in appropriate units. Because what is 

measured influences what we are able to perceive and 

respond to, in the case of Nr, it is especially critical to 

measure total amounts and different chemical forms at 

regular intervals over time. (Finding #13 – also pertains 

to study objective 4)
80

Specific Recommendations:

n The Committee recommends that EPA routinely and 

consistently account for the presence of Nr in the 

environment in forms appropriate to the medium 

in which they occur (air, land, and water) and that 

accounting documents be produced and published 

periodically (for example, in a fashion similar to 

National Atmospheric Deposition Program summary 

reports). The Committee understands that such an 

undertaking will require substantial resources, and 

encourages the Agency to develop and strengthen 

partnerships with appropriate federal and state agencies 

and private sector organizations with parallel interests 

in advancing the necessary underlying science of Nr 

creation, transport and transformation, impacts, and 

management. (Recommendation #13 – also pertains to 

study objective 4)

n EPA should work with USDA and universities 

to improve understanding and prediction of how 

expansion of biofuel production, as mandated by the 

2007 EISA, will affect Nr inputs and outputs from 

agriculture and livestock systems. Rapid expansion 

of biofuel production has the potential to increase N 

fertilizer use through expansion of corn production area 

and associated N fertilizer inputs, and from extending 

cultivation of cellulosic materials that will also need 

N inputs. Current models and understanding are not 

adequate to guide policy on how to minimize impact of 

biofuel expansion on environmental concerns related to 

Nr. (Recommendation #4)

n EPA should join with USDA, DOE, and universities 

in efforts to ensure that the N budgets of terrestrial 

systems are properly quantified and that the magnitudes 

of at least the major loss vectors are known. 

(Recommendation #9 – also pertains to study objectives 

2 and 4)

n EPA, USDA, DOE, and universities should work 

together to ensure that denitrification in soils and 

aquatic systems is properly quantified, by funding 

appropriate research. (Recommendation #10 – also 

pertains to study objective 4)

Study Objective #2: Evaluate the contribution an 

integrated N management strategy could make 

to environmental protection.

In general, the Committee finds that effective 

management of Nr in the environment must recognize 

the existence of tradeoffs across a number of impact 

categories involving the cycling of nitrogen and other 

elements. In addition, an integrated multi-media approach 

to monitoring Nr is needed. 

The Committee recommends that EPA: 

1. Develop a uniform assessment and management 

framework that considers the effects of Nr loading 

over a range of scales reflecting ecosystem, watershed, 

and regional levels. The framework should include 

all inputs related to atmospheric and riverine delivery 

of Nr to estuaries, their comprehensive effects on 

marine eutrophication dynamics and their potential for 

management 

2. Examine the full range of traditional and ecosystem 

response categories, including economic and 

ecosystem services, as a basis for expressing Nr 

impacts in the environment, and for building better 

understanding and support for integrated management 

efforts.

Specific Findings:

n There has been a growing recognition of eutrophication 

as a serious problem in aquatic systems (NRC, 2000). 

The last comprehensive National Coastal Condition 

Report was published in 2004 (EPA, 2004) and 

included an overall rating of “fair” for estuaries, 

including the Great Lakes, based on evaluation of 

over 2000 sites. The water quality index, which 

incorporates nutrient effects primarily as chlorophyll-a

and dissolved oxygen impacts, was also rated “fair” 

nationally. Forty percent of the sites were rated “good” 

for overall water quality, while 11% were “poor” and 

49% “fair.” (Finding #11)

n The Committee finds that reliance on only one 

approach for categorizing the measurement of Nr is 

unlikely to result in the desired outcome of translating 

N-induced degradation into the level of understanding 

needed to develop support for implementing effective 

Nr management strategies. (Finding #14)

n Effective management of Nr in the environment must 

recognize the existence of tradeoffs across impact 

categories involving the cycling of other elements, 

particularly carbon and phosphorus. (Finding #18)

n The Committee has determined that an integrated 

approach to monitoring that includes multiple media 

(air, land, and water) components and considers a 

suite of environmental and human concerns related 

to reactive nitrogen in the environment (e.g., Nr 

effects, climate change, human health) is needed. 

Some of the phenomena that presented in this 

report need more definition and verification. More 

importantly, however, as controls are brought to bear 

on Nr, improvements need to be measured to verify 

and validate successful management strategies. If 

the desired improvements are not realized as shown 

by the collected data, corrective measures will be 

required. Such an adaptive approach acknowledges 

the likelihood that management programs will be 

altered as scientific and management understanding 

improve. (Finding #20 – also pertains to study 

objective 3)

Specific Recommendations:

n The Committee recommends that EPA develop a 

uniform assessment and management framework that 

considers the effects of Nr loading over a range of 
81

scales reflecting ecosystem, watershed, and regional 

levels. The framework should include all inputs related 

to atmospheric and riverine delivery of Nr to estuaries, 

their comprehensive effects on marine eutrophication 

dynamics, and their potential for management. 

(Recommendation #11)

n It is recommended that the EPA examine the full range 

of traditional and ecosystem response categories, 

including economic and ecosystem services, as a basis 

for expressing Nr impacts in the environment, and for 

building better understanding and support for integrated 

management efforts. (Recommendation #14) 

n The Committee recommends that the integrated 

strategies for Nr management outlined in this report be 

developed in cognizance of the tradeoffs associated with 

reactive nitrogen in the environment, consistent with the 

systems approach of overarching recommendations 2 

and 3. Specific actions should include:

• Establishing a framework for the integrated 

management of carbon and reactive nitrogen

• Implementing a research program that addresses the 

impacts of tradeoffs associated with management 

strategies for carbon, reactive nitrogen, and other 

contaminants of concern

• Implementing a research and monitoring program 

aimed at developing an understanding of the 

combined impacts of different nitrogen management 

strategies on the interchange of reactive nitrogen 

across environmental media. (Recommendation #18)

n In cooperation with the Department of Agriculture, 

U.S. Army Corps of Engineers, U.S. Fish and 

Wildlife Service, and the and the Federal Emergency 

Management Agency, the EPA should develop 

programs to encourage wetland restoration and 

creation, with strategic placement of these wetlands 

where reactive nitrogen is highest in ditches, streams, 

and rivers. The Agency should also address the 

means of financing, governance, monitoring, and 

verification. Such programs might be modeled on the 

Conservation Reserve Program or extant water quality 

and environmental trading programs, but need not be 

limited to current practices. (Recommendation #15e – 

also pertains to study objective 3)

n The Committee recommends that EPA initiate 

discussions and take action to develop a national, 

multi-media monitoring program that monitors sources, 

transport and transition, effects using indicators where 

possible, and sinks of Nr in keeping with the nitrogen 

cascade concept. This comprehensive program should 

build upon existing EPA and state initiatives, as well as 

monitoring networks already underway in other federal 

agencies such as the U.S. Geological Survey programs 

and the NADP effort. (Recommendation #20 – also 

pertains to study objective 3)

Study Objective #3: Identify additional risk 

management options for EPA’s consideration.

In general, the Committee finds that a number of risk 

management actions should be considered to reduce Nr 

loading and transfer to the environment. The Committee 

recommends risk management actions that include farmlevel improvements in manure management, actions to 

reduce atmospheric emissions of Nr, and interventions to 

control Nr in water management programs. 

Specific Findings:

n Farm-level improvements in manure management can 

substantially reduce Nr load and transfer. There are 

currently very few incentives or regulations to decrease 

these transfer and loads, despite the existence of 

management options to mitigate. (Finding #6)

n Scientific uncertainty about the origins, transport, 

chemistry, sinks, and export of Nr remains high, but 

evidence is strong that atmospheric deposition of Nr 

to the earth’s surface, as well as emissions from the 

surface to the atmosphere, contribute substantially to 

environmental and health problems. Nitrogen dioxide, 

NO2, is often a small component of NOy, the total 

of oxidized nitrogen in the atmosphere. The current 

NAAQS for NO2, as an indicator of the criteria pollutant 

“oxides of nitrogen,” is inadequate to protect health 

and welfare. NOy should be considered seriously as a 

supplement or replacement for the NO2 standard and in 

monitoring. Atmospheric emissions and concentrations 

of Nr from agricultural practices (primarily in the form 

of NH3) have not been well monitored, but NH4

+ ion 

concentration and wet deposition (as determined by 

NADP and NTN) appear to be increasing, suggesting 

that NH3 emissions are increasing. Both wet and dry 

deposition contribute substantially to NHx removal, 

but only wet deposition is known with much scientific 

certainty. Thus, consideration should be given to adding 

these chemically reduced and organic forms of Nr to the 

list of criteria pollutants. (Finding #8)

n Meeting Nr management goals for estuaries, when a 

balance should be struck between economic, societal, 

and environmental needs, seems unlikely under current 

federal law. Enforceable authorities over nonpoint 

source, stormwater, air (in terms of critical loads), 

and land use are not adequate to support necessary Nr 

controls. Funding programs are presently inadequate 

to meet existing pollution control needs. Furthermore, 

new technologies and management approaches are 

required to meet ambitious Nr control needs aimed at 

restoring national water quality. (Finding #12 – also 

pertains to study objective 4)

n Intervention to control Nr under most water 

management programs generally occurs in three ways: 

• Prevention or source controls

• Physical, chemical or biological “dead ending” or 

storage within landscape compartments where it 
82

is rendered less harmful (e.g., long-term storage 

in soils or vegetation; denitrification, primarily in 

wetlands; reuse)

• Treatment using engineered systems such as 

wastewater treatment plants or BMPs for stormwater 

and nonpoint source runoff 

While most management programs focus on the 

third (treatment) approach, there are opportunities for 

combining the three that can be more effective and cost 

less. (Finding #15)

n The Committee finds that there have been persistent 

increases in the amounts of Nr that have been emitted 

into and retained within various ecosystems, affecting 

their functioning. Unless this trend is reversed, it 

will become increasingly difficult for many of these 

ecosystems to provide the services upon which human 

well-being depends. The Committee believes that there 

is a need to regulate certain forms of Nr to address 

specific problems related to excess Nr, and we believe 

that the best approach for an overall management 

strategy is the concept of defining acceptable total 

Nr critical loads for a given environmental system. 

(Finding #16 – also pertains to study objective 4)

n Current EPA policy (40 CFR Part 51, Clean Air Fine 

Particle Implementation Rule) discourages states from 

controlling ammonia emissions as part of their plan 

for reducing PM2.5 concentrations. In this rulemaking 

(Federal Register 72(79): 20586-20667), EPA has 

stated that “ammonia reductions may be effective 

and appropriate for reducing PM2.5 concentrations in 

selected locations, but in other locations such reductions 

may lead to minimal reductions in PM2.5 concentrations 

and increased atmospheric acidity.” Ammonia is a 

substantial component of PM2.5 in most polluted areas 

of the U.S. at most times. While it is true that reducing 

NH3 emissions might increase the acidity of aerosols 

and precipitation, the net effect of NH3 on aquatic and 

terrestrial ecosystems is to increase acidity. After being 

deposited onto the earth’s surface, NH4

+ is, under most 

circumstances, quickly nitrified, increasing the acidity 

of soils and waters. The Committee is unaware of any 

evidence that NH3 reduces the toxicity of atmospheric 

aerosols or that high concentrations of NH3 occur 

naturally over any substantial area of the U.S. It has 

not yet been established which components of PM 

have substantive effects on human health but the total 

concentration of PM2.5 correlates with morbidity and 

mortality, and NH3 contributes to PM2.5. The visibility, 

degradation, and other adverse effects associated with 

PM2.5 are related to aerosol surface area or mass where 

NH4

+ certainly plays a role. (Finding #17)

Specific Recommendations:

n A policy, regulatory, and incentive framework is 

needed and should be developed to improve manure 

management to reduce Nr load and ammonia 

transfer, taking into account phosphorus load issues. 

(Recommendation #6)

n EPA should re-examine the criteria pollutant “oxides of 

nitrogen” and the indicator species, NO2, and consider 

adding chemically reactive nitrogen as a criteria 

pollutant and NHx and NOy as indicators to supplement 

the NO2 National Ambient Air Quality Standards. 

(Recommendation #8a)

n The Committee recommends that monitoring of NHx

and NOy begin as soon as possible to supplement 

the existing network of NO2 compliance monitors. 

(Recommendation #8b) 

n The Committee recommends that EPA reevaluate 

water quality management approaches, tools, and 

authorities to ensure Nr management goals are 

attainable, enforceable, and the most cost-effective 

available. Monitoring and research programs should 

be adapted as necessary to ensure they are responsive 

to problem definition and resolution, particularly 

in the development and enhancement of nitrogen 

removal technologies and best management practices, 

and continue to build our level of understanding 

and increase our ability to meet management goals. 

(Recommendation #12 – also pertains to study 

objective 4)

n To better address Nr runoff and discharges from the 

peopled landscape, the Committee recommends that 

EPA:

• Evaluate the suite of regulatory and non-regulatory 

tools used to manage Nr in populated areas from 

nonpoint sources, stormwater and domestic sewage, 

and industrial wastewater treatment facilities, 

including goal-setting through water quality standards 

and criteria; and

• Determine the most effective regulatory and 

voluntary mechanisms to apply to each source 

type with special attention to the need to regulate 

nonpoint source and related land use practices. 

(Recommendation #15a)

• Review current regulatory practices for point sources, 

including both wastewater treatment plants and 

stormwater, to determine adequacy and capacity 

towards meeting national Nr management goals; and

• Consider technology limitations, multiple pollutant 

benefits, and funding mechanisms as well as potential 

impacts on climate change from energy use and 

greenhouse gas emissions, including nitrous oxide. 

(Recommendation #15b)

• Set Nr management goals on a regional/local basis, 

as appropriate, to ensure most effective use of limited 

management dollars; and

• Fully consider “green” management practices such as 

low-impact development and conservation measures 
83

that preserve or re-establish Nr removing features to 

the landscape as part of an integrated management 

strategy along with traditional engineered best 

management practices. (Recommendation #15c)

n The Committee recommends that the Agency work 

toward adopting the critical-loads approach concept 

in determining thresholds for effects of excess Nr on 

terrestrial and aquatic ecosystems. In carrying out 

this recommendation, the Committee recognizes that 

in many cases it will be necessary for the Agency to 

enter into new types of research, policy, and regulatory 

agreements with other federal, state, and tribal 

units based on cooperative, adaptive, and systemic 

approaches that derive from a common understanding 

of the nitrogen cascade. (Recommendation #16 – also 

pertains to study objective 4)

n The Committee recommends that the EPA presumption 

that NH3 is not a PM2.5 precursor should be reversed 

and states should be encouraged to address NH3 as a 

harmful PM2.5 precursor. (Recommendation #17)

Study Objective #4: Make recommendations to 

EPA concerning improvements in Nr research to 

support risk reduction.

The Committee has recommended research in 

key areas to support risk reduction. The Committee’s 

recommendations include research to advance the 

understanding of: the quantity and fate of Nr applied 

to major crops; how to accelerate crop yields while 

increasing N fertilizer uptake efficiency; agricultural 

emissions of forms of Nr; atmospheric deposition of 

Nr; and the potential for amplification of Nr-related 

climate impacts.

Specific Findings:

n Crop agriculture receives 60% of U.S. annual new 

Nr inputs from anthropogenic sources (9.8 Tg from 

N fertilizer, 7.7 Tg from crop BNF versus 29 Tg of N 

anthropogenically introduced into the U.S. environment 

per year ) and accounts for 58% (7.6 Tg) of total U.S. 

Nr losses from terrestrial systems to air and aquatic 

ecosystems, yet current monitoring of fertilizer use 

statistics by federal agencies is inadequate to accurately 

track trends in quantities and fate of N applied to major 

crops and the geospatial pattern by major watersheds. 

(Finding #1)

n Nr inputs to crop systems are critical to sustain crop 

productivity and soil quality. Moreover, given limited 

land and water resources, global population growth, 

and rapid economic development in the world’s most 

populous countries, the challenge is to accelerate 

increases in crop yields on existing farm land while 

also achieving a substantial increase in N fertilizer 

uptake efficiency. This process is called “ecological 

intensification” because it recognizes the need to 

meet future food, feed, fiber, and energy demand of 

a growing human population while also protecting 

environmental quality and ecosystem services for future 

generations (Cassman, 1999). More diverse cropping 

systems with decreased Nr fertilizer input may also 

provide an option on a large scale if the decrease in 

Nr losses per unit of crop production in these diverse 

systems can be achieved without a decrease in total 

food production, which would trigger indirect land 

use change to replace the lost production and negate 

the benefits. However, crop cultivars and agronomic 

practices are changing rapidly, which changes N 

requirements, but current efforts in research, extension, 

and conservation programs on N management within 

these rapidly evolving systems are not adequate to 

meet the challenge of providing better information to 

increase NFUE. (Finding #2)

n Nitrous oxide emissions from the Nr inputs to 

cropland from fertilizer, manure, and legume 

fixation represent a large proportion of agriculture’s 

contribution to greenhouse gas emissions, and 

the importance of this source of anthropogenic 

greenhouse gas will likely increase unless NFUE 

is markedly improved in crop production systems. 

Despite its importance, there is considerable 

uncertainty in the estimates of nitrous oxide emissions 

from fertilizer, and research should focus on reducing 

this uncertainty. (Finding #3)

n There are no nationwide monitoring networks in the 

U.S. to quantify agricultural emissions of greenhouse 

gases, NO, N2O, reduced sulfur compounds, VOCs, 

and NH3. In contrast, there is a large network in place 

to assess the changes in the chemical climate of the 

U.S. associated with fossil fuel energy production, i.e., 

the National Atmospheric Deposition Program/National 

Trends Network (NADP/NTN) which has been 

monitoring the wet deposition of sulfate (SO4

2-), NO3-

, 

and NH4

+ since 1978. (Finding #5)

n Synthetic N fertilizer application to urban gardens 

and lawns amounts to approximately 10% of the total 

annual synthetic N fertilizer used in the U.S. Even 

though this N represents a substantial portion of total 

N fertilizer use, the efficiency with which it is used 

receives relatively little attention. (Finding #7)

n The biogeochemical cycle of Nr is linked to climate 

in profound but nonlinear ways that are, at present, 

difficult to predict. Nevertheless, the potential for 

significant amplification of Nr-related impacts is 

substantial, and should be examined in more complete 

detail. (Finding #19)

Specific Recommendations:

n The Committee recommends increasing the specificity 

and regularity of data acquisition for fertilizer 

application to major agricultural crops in terms of 

timing and sufficiently small application scale (as 

well as for urban residential and recreational turf) by 

county (or watershed) to better inform decision-making 
84

about policies and mitigation options for reducing Nr 

load in these systems, and to facilitate monitoring and 

evaluation of impact from implemented policies and 

mitigation efforts. (Recommendation #1)

n To obtain information on Nr inputs and crop 

productivity the Committee recommends that:

• Data on NFUE and N mass balance, based on 

direct measurements from production-scale fields, 

be generated for the major crops to identify which 

cropping systems and regions are of greatest concern 

with regard to mitigation of Nr load, and to better 

focus research investments, policy development, 

and prioritization of risk mitigation strategies. 

(Recommendation #2a)

• Efforts at USDA and universities be promoted to: 

(1) investigate means to increase the rate of gain in 

crop yields on existing farm land while increasing 

N fertilizer uptake efficiency, and (2) explore the 

potential for more diverse cropping systems with 

lower N fertilizer input requirements to the extent 

that large-scale adoption of such systems would not 

cause indirect land use change. (Recommendation 

#2b)

• EPA work closely with the U.S. Department of 

Agriculture (USDA), Department of Energy (DOE), 

the National Science Foundation (NSF), and 

universities to help identify research and education 

priorities to support more efficient use and better 

mitigation of Nr applied to agricultural systems. 

(Recommendation #2c)

n The Committee recommends that EPA ensure that the 

uncertainty in estimates of nitrous oxide emissions 

from crop agriculture be greatly reduced through the 

conduct of EPA research and through coordination of 

research efforts more generally with other agencies 

such as USDA, DOE, NSF, and with research 

conducted at universities. (Recommendation #3)

n The status and trends of gases and particulate matter 

emitted from agricultural emissions (e.g., NO3

-

 and 

NH4

+) should be monitored and assessed utilizing a 

nationwide network of monitoring stations. EPA should 

coordinate and inform its regulatory monitoring and 

management of reactive nitrogen with the multiple 

efforts of all agencies including those of the U.S. 

Department of Agriculture and NSF-supported efforts 

such as the National Ecological Observatory Network 

(NEON) and the Long Term Ecological Research 

Network (LTER). (Recommendation #5)

n To ensure that urban fertilizer is used as efficiently as 

possible, the Committee recommends that EPA work 

with other agencies such as USDA as well as state and 

local extension organizations to coordinate research 

and promote awareness of the issue. (Recommendation 

#7a)

n Through outreach and education, supported by 

research, improved turf management practices should 

be promoted, including improved fertilizer application 

and formulation technologies and maintenance 

techniques that minimize supplemental Nr needs and 

losses, use of alternative turf varieties that require less 

fertilization, alternative ground covers in place of turf, 

and use of naturalistic landscaping that focuses on 

native species. (Recommendation #7b)

n EPA should pursue the longer term goal of monitoring 

individual components of Nr, such as NO2 (with 

specificity), NO, PAN, and HNO3, and other inorganic 

and reduced forms, as well as support the development 

of new measurement and monitoring methods. 

(Recommendation #8c)

n The scope and spatial coverage of the Nr concentration 

and flux monitoring networks (such as the National 

Atmospheric Deposition Program and the Clean Air 

Status and Trends Network) should be increased, and 

an oversight review panel should be appointed for these 

two networks. (Recommendation # 8d)

n EPA in, coordination with other federal agencies, 

should pursue research goals including:

• Measurements of deposition directly both at the 

CASTNET sites and in nearby locations with nonuniform surfaces such as forest edges

• Improved measurements and models of convective 

venting of the planetary boundary layer (the lowest 

layer of the atmosphere) and of long range transport

• Improved analytical techniques and observations 

of atmospheric organic N compounds in vapor, 

particulate, and aqueous phases

• Increased quality and spatial coverage of 

measurements of the NH3 flux to the atmosphere 

from major sources especially agricultural practices

• Improved measurement techniques for, and numerical 

models of, NOy and NHx species (especially 

with regard to chemical transformations, surface 

deposition and off-shore export and linked ocean-landatmosphere models of Nr). (Recommendation #8e)

n Research should be conducted on: best management 

practices that are effective in controlling Nr, especially 

for nonpoint and stormwater sources (including 

land and landscape feature preservation); setting 

Nr management targets that realistically reflect 

these management and preservation capacities; and 

constructing a decision framework to assess and 

determine implementation actions consistent with 

management goals. (Recommendation 15d)

n The EPA should support cross-disciplinary and 

multiagency research on the interactions of 

climate and Nr. To determine the interactions of 

global biogeochemical Nr cycles and climate, the 
85

Committee suggests that EPA follow a series of steps 

such as:

1. Select several likely scenarios for global climate from 

the IPCC report for the year 2050

2. Down-scale statistics or nest regional climate models 

within each of these global scenarios to generate 

meteorological and chemical fields (e.g., temperature, 

relative humidity, winds, precipitation, CO2) for a few 

years around 2050

3. Run several independent biogeochemical Nr models 

(earth system models that include air/water/land) for 

North America for these years with current Nr and 

emissions and application rates

4. Rerun models with decreased Nr emissions/application 

to evaluate strategies for controlling impacts such as 

those described in this report. (Recommendation #19)

6.5 Conclusions and Observations

Nitrogen is not only an essential resource for humans, 

for example its use in food production, but also a 

byproduct of essential processes such as combustion. 

The increasing amounts of nitrogen that humans capture 

through the Haber-Bosch process for useful purposes 

and the increasing amounts of nitrogen released from 

combustion also add increasing amounts of excess 

reactive nitrogen (Nr) to the environment. In the United 

States, the production of crops receives 60% of the 

new Nr inputs from anthropogenic sources, but also 

accounts for almost the same proportion of total Nr losses 

from terrestrial systems to air and aquatic ecosystems. 

Nitrogen fertilizer use efficiency by crops seldom reaches 

50% and may be 33% or lower. Fossil fuel combustion is 

another major contributor of excess Nr in the production 

of essential energy. However, the excess Nr that flows 

from these activities is not benign. There are serious 

negative impacts from excess Nr on both human health 

and the environment. Table 1 gives a range of examples 

of these.

The nature of the problem

Dealing with excessive reactive nitrogen is an 

extraordinarily complex issue. Part of this relates to the 

nature of nitrogen and its ability to change its form and 

flow through different media, as evidenced conceptually 

by the nitrogen cascade. Nitrogen’s transformative nature 

only increases the difficulty of dealing with its negative 

aspects. Further, unlike the linear problems that society 

has been more accustomed to dealing with in the past, 

excessive Nr in the environment is in a class of problems 

sometimes characterized as “wicked” (Batie 2008, 

Kreuter 2004). For a “wicked” problem;

• There is not universal agreement on what the problem 

is – different stakeholders define it differently

• There is no defined end solution, the end will be 

assessed as “better” or worse”

• The problem changes over time

• There is no clear stopping rule – stakeholders, 

political forces and resource availability will make 

that determination on the basis of “judgments”

• The problem is associated with high uncertainty of 

both components and outcomes

• Values and societal goals are not necessarily shared 

by those defining the problem or those attempting to 

make the problem better

This is not to say that society does not try to deal with 

such problems (healthcare, environmental degradation, 

water resource management, food safety, etc.). What it 

does say is that different approaches are needed to deal 

with such problems as compared with better defined 

problems that are amenable to disciplinary linear science 

where experts can define the problem and its endpoint. 

The Integrated Nitrogen Committee has provided a 

number of findings and recommendations with respect to 

the problems caused by excess Nr. Some of these respond 

to specific defined science based concerns. Others relate 

to broader concerns the Committee raised in dealing with 

the overall Nr dilemma. The following points synthesize 

some of the important lessons learned from the more 

than four years of study on this issue by the Integrated 

Nitrogen Committee.

Recognizing the problem and building 

consensus

It is critically important that the problems caused 

by integrated nitrogen be widely recognized. Many 

recognize that we add nitrogen to the environment for 

specific useful purposes. Fewer recognize that there are 

both direct and indirect impacts from this that damage 

both the environment and human health. Until there is 

general recognition that there is a problem and we need 

to deal with the negative externalities of excess reactive 

nitrogen in our environment, there will be no willingness 

to tackle this issue. Education, communication and 

outreach are critically important to engender in the 

public sufficient will to tackle this widespread problem. 

Education, communication and outreach will be critical 

to the formation of a common definition of the problem, 

an essential step once it is recognized. Following that, 

there will have to be some degree of public consensus 

for actions that will effectively reduce excess reactive 

nitrogen in the environment. These steps will not be 

possible unless there is a process of public consensus 

building with respect to Nr. The first essential step in 

trying to deal with a “wicked” problem is getting some 

measure of agreement across different participants and 

stakeholders about the problem itself. 

The importance of the nitrogen cascade

Understanding the problem will require recognition 

of the nature of Nr. In this case, the Integrated Nitrogen 

Committee found the concept of the nitrogen cascade 
86

(pictured in the Executive Summary and Chapter 1) 

to be an essential guide to approaching the problem. 

This conceptual framework traces the flows of new 

nitrogen through atmospheric, terrestrial, and aquatic 

environmental systems where Nr is received, stored 

and passed on in one form or another. The Committee 

initially spent much of its effort on understanding 

these flows and then determining as best as possible 

the magnitudes of the flows and sinks through these 

systems. This is the critical baseline information 

needed to understand the problem, and many of the 

recommendations in this report relate to gaining a better 

understanding of this phenomenon. Unless it is well 

understand how Nr flows through these systems, what 

the Nr sinks are in these system components, and what 

the magnitudes are of both the sinks and flows, any 

attempts to better control Nr may well be ineffective. 

Understanding and being able to quantify the nitrogen 

cascade will allow the identification of the major excess 

Nr contributors, an understanding of the different forms 

of Nr at various points along the cascade, and enable 

the determination of more effective interdiction points. 

Without this basic understanding of the flows of Nr, their 

nature, and their magnitude, attempts to deal with excess 

Nr will have high uncertainty. Chapter 2 summarizes 

much of this critical information that the Committee was 

able to determine and assemble from other sources.

Integrated approaches are essential

Given what we know about the way Nr behaves, 

efforts to deal with excess Nr must be organized in a way 

that reflects the nature of the problem. Unfortunately, 

many of our approaches are narrowly disciplinary 

focused, and our policy and regulatory institutions are 

often focused on one or another media where excess Nr 

may temporarily reside or on a sector that contributes 

excess Nr. The regulatory structure that has evolved for 

problems affecting human health and the environment 

has been specifically narrow, following policymakers 

focus on such things as clean water or clean air. These 

silos have to be broken down if excessive Nr is to be 

dealt with effectively. Current efforts by EPA and other 

agencies to encourage more integration across these 

silos, and include other institutions and stakeholders, 

are absolutely essential. Research efforts to better 

understand excess Nr and better mitigate its negative 

impacts must be trans-disciplinary. What will also be 

critical is greatly increased collaboration and cooperation 

among and between agencies and interested stakeholders. 

Several of the Committee’s recommendations address 

this issue, calling for formal mechanisms to encourage 

this to take place. The Nr problem has boundaries much 

broader than the boundaries of the institutions that we 

have relied on in the past to protect human health and 

the environment. We are not going to be effective if we 

do not both expand these boundaries and adopt a broad 

multi-institutional reach.

Essential monitoring and research

In the effort to understand and quantify the nitrogen 

cascade the Committee became aware of areas of needed 

research and monitoring. Much of this is essential for 

improving our knowledge of what goes on in the cascade. 

It is also essential for benchmarking current Nr flows and 

sinks and for targeting where actions are best taken to 

reduce excess Nr. There is not sufficient information and 

understanding of these flows and sinks to allow maximum 

benefit from utilizing the full power of the cascade 

concept. The Committee has recommended improved 

monitoring and research to enhance our understanding 

in air, land and water environments. In some cases 

our knowledge has such wide margins of error that we 

cannot identify or quantify important concentrations or 

flows sufficiently for necessary decisions. In other cases 

we need much better understanding of the efficacy of 

actions that might be taken to control Nr. In some cases 

we need to know more about the indirect impacts of Nr 

as well as the indirect impacts of measures to control 

Nr. Monitoring is both an essential part of the research 

needs as well as being a critical guide to what we face 

and whether our efforts to better control Nr are being 

effective. Such environmental monitoring is often not 

considered to be critically important. It is critically 

important for effective approaches to reducing excess Nr.

Where to begin

The approach taken in this report and its 

recommendations should result in helping enable the 

control of excess Nr to proceed on the basis of starting 

where it is most technically and economically effective. 

Some of this may be low hanging fruit, or particular 

niches where institutions and stakeholders already have 

a common purpose. Part of the value of the nitrogen 

cascade is that its characterization of Nr sinks and 

flows allows the comparison of alternative points of 

entry and interdiction. Further monitoring and research 

should also allow the comparison of different modes of 

interdiction. There will be choices to be made between 

preventing increased Nr at the source, stabilizing it, or 

treating the medium involved to remove it. Additional 

research and actual experience in such efforts, coupled 

with monitoring, will be essential for making these 

choices. One product of the cascade exercise is a better 

understanding of what and where the major contributors 

of Nr are. The big sources of Nr have to be addressed 

if there is to be meaningful impact on reducing excess 

Nr. Part of the decision about where to begin will relate 

to the efficacy of measures to reduce excess Nr, and our 

knowledge is by no means complete on this. Beyond 

the technical considerations of effectiveness, there will 

have to be policy decisions about the trade-offs and 

interdependencies between approaches such as market 

mechanisms, regulation, incentives, and voluntary actions. 

Market mechanisms generally require regulations that 

are the basis for the creation of the market situation that 
87

makes such things as emission trading viable. Voluntary 

actions are more likely if there is some prospect of future 

required action. 

Metrics matter

What is measured is critical to determining the 

dimension of the problem, what to tackle, and whether 

progress is being made. Kilograms removed (or prevented 

at the source), the percentage of Nr removed from a 

particular cascade or sink, or dollars of damage avoided 

by the removal of Nr are all reasonable measures to use. 

Measuring the improvement in physical units of Nr can 

lead to very different source targets, control measures and 

estimations of success as compared to measuring dollars 

of damage. This relates to the dilemma presented by 

“wicked” problems in both the original problem definition 

and the determination of whether actions have made 

things better or worse. An example from the Chesapeake 

Bay is cited in the report that illustrates this. If the goal is 

less Nr in the bay waters, then all sources are important. 

If one is concerned about reducing the economic damage 

of Nr and uses dollars of damage, then atmospheric 

deposition becomes the primary focus because of the 

high value of health damages from atmospheric pollution 

related to Nr. 

Setting goals for action

Finally, nothing is going to be accomplished if goals 

are not set and efforts do not get underway. There are 

sufficient findings and recommendations presented by 

the Committee in addition to what we already know to 

enable multiple agencies to begin to reduce the excess 

Nr entering the environment. Some of the trade-offs 

between alternative approaches are also well known. 

The Committee suggests actions that might be taken by 

EPA or other management authorities to reduce nitrogen 

in the environment and a 25% reduction of excess Nr is 

suggested as attainable with current technology over the 

near term. Actions being suggested need testing, refining 

and require monitoring. A start, even on a pilot scale, in 

one portion of the cascade or sector will yield valuable 

information about the efficacy of the approaches used, 

further demonstrate the necessity for a multi-agency joint 

stakeholder approach, and help further define the problem 

and where it can initially be best addressed. This necessity 

of getting underway is one of the main recommendations 

of the National Research Council’s series of reports on 

improving water quality in the Mississippi (National 

Research Council, 2008b and 2009). These reports 

also emphasize the institutional arrangements that are 

necessary. These are also echoed in this report. Following 

on the National Research Council’s reports, the USDA 

Natural Resources Conservation Service has recently 

begun an effort that targets nutrient reductions in the 

Upper Mississippi. It is time to get more efforts underway 

with effective collaboration between public and private 

institutions and stakeholders.

Because this report addresses the needs of EPA’s 

research mission, there is substantial emphasis on gaps 

in knowledge about Nr and the research that needs to 

be done to fill these gaps. However, the report contains 

an extensive knowledge base about Nr which was a 

necessary precursor to addressing the objectives of the 

report. The report then identifies the problems posed 

by Nr, assesses the necessity of an integrated strategy 

to deal with these problems, identifies some risk 

management options for EPA’s consideration, and makes 

recommendations for improved research and monitoring 

to support risk reduction. This information is sufficient 

to allow initial determinations of where and how the Nr 

problem can be addressed effectively with positive results. 

As efforts progress, more will be learned and improved 

methods, targeting, and analysis can be applied to this 

truly wicked problem.
88
89

Review of Nr wet deposition

Substantial progress has been made in monitoring Nr 

wet deposition as is summarized in information provided 

by the National Atmospheric Deposition Program/

National Trends Network (NADP) established in 1979. 

This network monitors precipitation composition at over 

250 sites in the U.S. and its territories (http://nadp.sws.

uiuc.edu). Precipitation at each station is collected weekly 

according to well established and uniform procedures and 

sent to the Central Analytical Laboratory for analysis of 

acidity, NO3

-

, NH4

+, chloride, as well as the base cations 

calcium, magnesium, potassium and sodium. For greater 

temporal resolution, the Atmospheric Integrated Research 

Monitoring Network AIRMON, composed of seven 

sites, was formed in 1992 as part of the NADP program 

to study wet deposition composition and trends using 

samples collected daily. The same species are measured 

as in NADP. By interpolating among sites, NADP is able 

to estimate the wet deposition of NH4

+ (reduced N), and 

NO3

-

 (oxidized N) for the 48 contiguous states (Table A-1 

and Figure A-1). 

Although individual regions vary, the NADP data for 

the entire 48 states indicate an apparent decrease in NO3

-

wet deposition, but not in NH4

+ deposition (Table A-1 and 

Figure A-2). Ammonium wet deposition shows a weak 

increase, although the correlation coefficient is small. As 

NOx controls have become more effective, the role of 

reduced N appears to have grown in relative importance. 

The nitrate data appear to show a statistically significant 

trend and quantifying the response of deposition to a 

change in emissions would be useful to both the scientific 

and policy communities. A notable reduction in power 

plant NOx emissions occurred as the result of the NOx

State Implementation Plan (SIP) call (McClenny et al., 

2002; Gilliland et al., 2008; Bloomer et al., 2009). EPA 

should pursue a rigorous analysis of the emissions and 

deposition data, including identifying monitors and 

methods that are consistent from the beginning to the end 

of the record, as indicated in Recommendation 8. 

How is Nr deposition related to emissions?

The relationship between emissions of Nr and 

observed deposition is critical for understanding the 

efficacy of abatement strategies as well as for partitioning 

local and large-scale effects of emissions. Only a few 

studies covering several individual sites have sufficient 

monitoring consistency and duration to determine 

rigorously long-term trends in NO3

-

 and NH4

+ and their 

relationship to emissions, and here we consider several 

Appendix A:

Nitrogen Deposition from the Atmosphere to 

the Earth’s Surface

Table A-1: Annual wet deposition of reduced 

(NH4

+), oxidized (NO3-

), and total N to the 48 

contiguous states 

Source: NADP/National Trends Network (NTN) http://

nadp.sws.uiuc.edu

Figure A-1: Percent contribution of oxidized 

(NO3

-

) and reduced (NH4

+) nitrogen wet deposition from 1994 to 2006. 

As emissions of NOx have decreased, the relative 

importance of NHx has increased (data from National 

Atmospheric Deposition Program, 2010). Atmospheric wet N deposition USA

NADP/NTN deposition estimates
90

examples (Kelly et al., 2002; Butler et al., 2005; Likens et 

al., 2005). These sites tend to be in the eastern U.S. where 

monitoring is more concentrated and has a longer history 

and where upwind sources and downwind receptors are 

relatively well known. Examination of these studies 

reveals that concentrations of gaseous and particulate 

N species in the atmosphere, as well as the Nr content 

of precipitation over the eastern U.S., shows significant 

decreases. Correlation with regional emissions is stronger 

than with local emissions, in keeping with the secondary 

nature of the major compounds – NO3

-

 and NH4

+. 

Decreases in NH4

+ concentration and wet deposition are 

attributed to decreases in SO4

2– concentrations, meaning 

that more of the reduced Nr remains in the gas phase. For 

the period 1965 to 2000, NO3

-

 levels in bulk deposition 

correlate well with reported NOx emissions. For shorter 

and earlier time periods the correlation is weaker, and the 

authors attribute this to changes in the EPA’s methods of 

measuring and reporting emissions; they find evidence of 

continued errors in emissions from vehicles. Decreases in 

deposition will probably not be linearly proportional to 

decreases in emissions; for example a 50% reduction in 

NOx emissions is likely to produce a reduction of about 

35% in concentration and deposition of nitrate. 

The relationship between chemically reduced N 

emissions and deposition is more complex. The maps of 

ammonium deposition (Figure A-3) show that maxima 

occur near or downwind of major agricultural centers 

where emissions should be high. The full extent of the 

deposition record (see http://nadp.sws.uiuc.edu) shows the 

large intensification of NH4

+ wet deposition in selected 

areas. The southeastern U.S., particularly North Carolina, 

has seen a long-term rise (Aneja et al., 2001; Aneja 

et al., 2003; Stephen and Aneja, 2008). The increase 

in deposition coincides with the increase in livestock 

production, but a swine population moratorium appears to 

have helped abate emissions (Stephen and Aneja, 2008). 

Ammonium deposition, kgN/ha/yr Nitrate deposition, kgN/ha/yr

Figure A-2: Trend in reported wet deposition of 

NH4

+ and NO3-

 for the 48 contiguous states

Note the sampling methods and locations have not been 

tested for temporal or spatial bias (data from National 

Atmospheric Deposition Program, 2010).

Figure A-3: Annual NH4

+, NO3-

, and total 

inorganic N deposition for the year 2007 showing 

spatial patterns of deposition 

Source: National Atmospheric Deposition Program, 2010.
91

Concentrations of aerosol NH4

+ have decreased in many 

parts of the country, and this may appear to contradict 

the trend in wet deposition, but a decrease in condensed 

phase NH4

+ will be accompanied by an increase in 

vapor phase NH3 if SO4

2– and NO3-

 concentrations 

decrease; see http://vista.cira.colostate.edu/improve/. This 

potentially misleading information highlights the need for 

measurements of speciated NHx (Sutton et al., 2003). 

Review of dry deposition observations for the 

eastern United States

Monitoring dry deposition presents a greater 

challenge than monitoring wet deposition. The Clean 

Air Standards and Trends Network (CASTNET) and 

Atmospheric and Integrated Research Monitoring 

Network (AIRMON) were established to monitor 

chemical and meteorological variables to infer dry 

deposition in order to study the processes leading from 

emissions to atmospheric concentrations and through 

deposition to ecosystem effects. AIRMON dry deposition 

monitoring was discontinued in 2003. See www.epa.gov/

castnet/, www.arl.noaa.gov/research/programs/airmon.

html, and http://nadp.sws.uiuc.edu.

Recent reviews (Sickles and Shadwick, 2007a,b) 

analyze the seasonal and regional behavior of 

concentration and deposition of a variety of primary and 

secondary pollutants including Nr and investigated trends 

from 1990 to 2004 for the U.S. east of the Mississippi 

River. The investigators evaluated observations from 

more than 50 sites in the eastern states and concluded 

that for 2000-2004, the mean annual total measured N 

deposition for this area was 7.75 kg N per hectare per 

year (expressed as kg N/ha/yr); see Table A-2. This value 

includes vapor phase HNO3, particulate NO3

-

, and NH4

+; 

it does not include deposition of other oxidized species 

such as NOx and PAN, nor gas-phase reduced N species, 

most notably NH3. The measured deposition rates peak 

in spring and summer, but unaccounted for ammonia 

deposition is probably a substantial fraction of the total, 

and the true annual cycle remains uncertain. 

Estimated total N deposition to the eastern 

United States

CASTNET monitors HNO3 and NO3

-

, but not other 

members of the NOy family – notably NOx. Dennis 

(U.S. EPA, 2007d) estimated that the unmeasured NOy

species account for about 50% of the dry deposition of 

nitrates. Half of 1.88 (see Table A-2) is 0.94 kg N /ha/yr. 

Ammonia is also unmeasured by CASTNET, and model 

estimates (Mathur and Dennis, 2003) of NH3 indicate 

that dry deposition should account for 75% of wet NH4

+

deposition; 75% of 2.54 is 1.9 kg N /ha/yr. Adding these 

two values to the total from Table A-2 yields a reasonable 

estimate, within about ±50% absolute accuracy, of total 

deposition of about 10.6 kg N /ha/yr for the eastern U.S. 

Characteristics of N deposition to the eastern 

United States

Analysis of production of N2 and N2O via gas phase 

reaction is provided in Appendix E. Warmer temperatures 

are conducive to release of NH3 from soils and manure 

as well as from atmospheric particles, thus ammonia 

concentrations are typically highest in summer. Diffusion 

of gases is faster than diffusion of particles, and dry 

deposition of vapor-phase Nr is faster as well; for 

example the mean CASTNET reported HNO3 deposition 

velocity is 1.24 cm/s while that for particulate NO3

-

 is 

0.10 cm/s. In 2003 and 2004 substantial reductions in 

emissions from electricity-generating units (power plants) 

were implemented under the NOx State Implementation 

Plan (SIP) call. Many of these power plants are 

located along the Ohio River generally upwind of the 

measurement area. Significant reductions (p = 0.05) were 

found between the 1990-1994 and 2000-2004 periods 

(Sickles and Shadwick, 2007a). 

Uncertainty in measured deposition

Analysis of uncertainties in the deposition of Nr is 

challenging. The coefficient of variation for total regional 

N deposition for 2000-2004 is 23%, representing a 

minimal value of uncertainty. Concentrations of some of 

the NOy species are monitored, as is the wet deposition of 

major oxidized and reduced N species, but concentrations 

of ammonia and other Nr species are not monitored. The 

network for monitoring dry deposition is sparse and has 

not been evaluated for spatial bias. The monitors are 

located in flat areas with uniform surfaces – advective 

deposition into, for example, the edges of forests are 

estimated to contribute substantially to the uncertainty 

(Hicks, 2006). Other sources of error include the model 

Table A-2: Deposition of N to the eastern United 

States in units of kg N/ha/yr

Annual deposition

kg N/ha/yr

Dry NH4

+ 0.41

Wet NH4

+ 2.54

Dry HNO3 + NO3

- 1.88

Wet NO3

- 2.92

Total measured N Dep. 7.75

Est. dry other NOy 0.94

Est. dry NH3 1.90

Est. total NOy 5.74

Est. total NH3 + NH4

+ 4.85

Est. Grand Total 10.59

Data are from the U.S. CASTNET program for the period 

of 2000-2004. Monitored species for 34 sites east of the 

Mississippi include vapor-phase HNO3, particulate NO3

-

, 

and NH4

+; unmonitored are other oxidized species such 

as NOx and PAN and gas-phase reduced N species 

most notably NH3 (Sickles and Shadwick, 2007a). For 

an explanation of how deposition of unmeasured species 

was estimated see text.
92

used to convert weekly average concentrations and 

micrometeorological measurements into depositions. 

Precision can be determined from collocated sites and 

is estimated at 5% for nitrate and 15% for ammonium 

in precipitation (Nilles et al., 1994). The uncertainty in 

estimated dry deposition arises primarily from uncertainty 

in deposition velocities (Hicks et al., 1991; Brook et 

al., 1997) and can be as high as 40% for HNO3. Total 

uncertainty for deposition of Nr based on measurements is 

at least 25% and may be as high as 50%.

Deposition estimates from numerical models

The EPA Community Multiscale Air Quality model 

(CMAQ) was run for North America at 36 km resolution 

(R. Dennis et al., personal communication, January 2008). 

Simulation of Nr deposition is hampered by the lack of 

emissions information (especially for NH3), by the need 

to parameterize planetary boundary layer (PBL) dynamics 

and deep convection, as well as by simplified multiphase 

chemistry. This run of CMAQ did not account for NOx

emissions from marine vessels, and these amount to 

about 4% of the total NOx emissions in 2000. Calculated 

nitrogen deposition for the 48 contiguous states (Table 

A-3) was broadly consistent with direct measurements 

(Table A-2). CMAQ NOx emissions were 5.84 Tg N for 

the year 2002; of that, 2.74 Tg N were deposited. This 

suggests that ~50% was exported – a number somewhat 

higher than has been reported in the literature; this 

discrepancy is discussed below.

Ammonia emissions and ambient concentrations can 

be measured, but are not routinely monitored. For Nr, the 

CMAQ numerical simulation employed inverse modeling 

techniques – that is, NH3 emissions were derived from 

observed NH4

+ wet deposition (Gilliland et al., 2003; 

Mathur and Dennis, 2003; Gilliland et al., 2006). Model 

determinations of NH3 therefore do not provide an 

independent source of information on NH4

+ deposition. 

The three-year CMAQ run gives an indication of the 

spatial pattern of deposition (Figures A-4). For NHx, wet 

and dry are equally important, but for NOy, dry deposition 

accounts for about two-thirds of the total deposition, 

while wet deposition accounts for about one-third. While 

this does not hold for the eastern U.S. it is true for the 

U.S. as a whole; in arid southern California, for example, 

dry deposition of Nr dominates. Based on CMAQ, total 

NOy deposition is 2.79 times the wet deposition and total 

NHx deposition is 1.98 times the wet deposition. Using 

the data from Table A-1 for the average wet deposition for 

the period 2000- 2004, total deposition of oxidized N is 

4.36 kg N /ha/yr (2.79 × 1.56 = 4.36). The total deposition 

for reduced N is 3.17 kg N /ha /yr (1.98 × 1.60). The 

grand total (wet and dry oxidized and reduced) is then 

about 7.5 kg N /ha /yr. 

The model has highly simplified organic N deposition. 

Note that these values reflect emissions before the NOx

SIP-call, which resulted in substantial reductions in NOx

emissions from point sources over the eastern U.S.

For comparison purposes, a collection of chemical 

transport models (CTMs) (Dentener et al., 2006) yielded 

total (wet plus dry) deposition to the whole U.S. of about 

3.9 Tg N /yr oxidized Nr and 3.0 Tg N /yr ammoniacal 

N for current emissions. The fate of NOx is assumed to 

be primarily HNO3 or aerosol NO3

-

; organic N species 

are generally not modeled in detail. Because this analysis 

includes Alaska, a better estimate for NOx for the 48 

contiguous states is 4.6 Tg N /yr. The variance among 

models was about 30% (one standard deviation) for 

deposition fluxes in regions dominated by anthropogenic 

emissions. Globally, the calculations from the ensemble 

of 23 CTMs estimated that 36-51% of all NOy and NHx

emissions are deposited over the ocean. This load could be 

important to estuarine N loading estimates, as offshore N is 

carried inshore by currents or through advective processes.

Deposition estimates from mass balance

From estimated total emissions of Nr compounds 

and observed or simulated export, a reasonable estimate 

of rate of deposition can be obtained by mass balance 

– deposition equals emissions minus export. Although 

substantial uncertainty (about a factor of two) exists for 

the emissions of NH3, NOx release is reasonably well 

known. In general, advection in the boundary layer and 

lofting through convection followed by export at higher 

altitudes are the two main mechanisms that prevent 

removal of NOy and NHx by deposition to the surface of 

North America (Luke et al., 1992; Li et al., 2004). 

Experimental observations have been conducted over 

the eastern U.S. for more than two decades (Galloway 

et al., 1984; Galloway and Whelpdale, 1987; Luke and 

Dickerson, 1987; Galloway et al., 1988). Most recent 

estimates (Dickerson et al., 1995; Li et al., 2004; Parrish 

et al., 2004b; Hudman et al., 2007), agree that annually 

7-15% of the emitted NOx is exported in the lower to 

mid-troposphere. 

CTMs derived small export values – on the order of 

30% of the total NOx emitted into the lower atmosphere 

Table A-3: Results from CMAQ* for total 

deposition in 2002 to the 48 contiguous states of 

oxidized and reduced N

kg N/ha/yr Tg N/yr

Oxidized N 3.51 2.74

Reduced N 2.66 2.07

Total N Deposition 6.17 4.81

The CMAQ results were adapted from Schwede et 

al., (2009) The Watershed Deposition Tool: A tool for 

incorporating atmospheric deposition in water-quality 

analyses, htttp://www.epa.gov/amad/EcoExposure/

depositionMapping.html.
93

(Penner et al., 1991; Kasibhatla et al., 1993; Holland 

et al., 1997; Horowitz et al., 1998; Liang et al., 1998; 

Galloway et al., 2004; Park et al., 2004; Holland et al., 

2005; Doney et al., 2007). Reviewed publications using 

the mass balance approach have substantial uncertainty 

but indicate with some consistency that 25-35% of the 

NOy emitted over the U.S. is exported. 

Comparison of models and measurements of 

oxidized N deposition

Both ambient measurements and numerical models 

of NOy have reached a level of development to allow 

reasonable estimates of deposition. For reduced 

nitrogen, neither ambient concentrations nor emissions 

are known well enough to constrain models. Recent 

model estimates of the U.S. N budget are reasonably 

uniform in finding that about 25-35% of total NOx

emissions are exported. 

Results from CMAQ runs described above indicate 

that, of the NOx emitted over the continental U.S., 50% 

is deposited and 50% is exported. This is within the 

combined error bars of other studies, but well under the 

best estimate of 70% deposition. One possible source 

of this discrepancy is underestimation of deposition of 

organo-nitrogen compounds. The chemical mechanism 

used in CMAQ was highly simplified – only about 2-3% 

of the total Nr deposition can be attributed to organonitrogen compounds (R. Dennis personal communication, 

2008). Ammonia from fossil fuel combustion, while 

important locally, is probably a small component of 

national Nr deposition. 

Major sources of uncertainty in modeled and observed 

values include missing deposition terms and poorly 

constrained convective mass flux. As indicated above, 

convective mass flux (rapid vertical transport) is uncertain 

because most convective clouds are smaller than a grid 

box in a global model. There is evidence for nonlinearities 

in NO2 deposition velocities with greater transfer from the 

atmosphere to the surface at higher concentrations (Horii 

et al., 2004; 2006). 

Figure A-4: CMAQ annual average (wet plus dry and oxidized plus reduced) nitrogen deposition (in 

kg-N/ha/yr) across the United States. 

This is based on three years of differing meteorology – one dry, one wet, and one average precipitation year – across 

the Eastern United States 

Source: U.S. Environmental Protection Agency, 2007f
94

Emissions from Canada and Mexico can have a 

substantive impact on atmospheric Nr over the U.S. near 

major sources, such as downwind of industrial Ontario 

and major cities of San Diego, CA, and Tijuana, Mexico 

(Wang et al., 2009). While Nr is imported into the U.S. 

from these border countries, there is also export. The 

emissions from Canada and Mexico are each 10-15% of 

those of the U.S. and the bulk of the Mexican population 

is distant from the U.S. We expect the overall impact of 

neighboring countries to add about 10% uncertainty to the 

estimated Nr budget for the 48 contiguous states. 
95

Appendix B:

Sources and Cycling of Nr Input into 

Terrestrial Systems in the United States

Most of the new Nr introduced into terrestrial 

systems in the U.S. was used to produce food for human 

consumption and forage and feed for livestock and poultry 

(~17.7 Tg total with 9.7 Tg from synthetic fertilizer and 

~8 Tg from biological N fixation; Table B-1). In addition 

to new Nr and Nr that was recycled from livestock and 

human excreta, crop production releases Nr that was 

stored in soil organic matter (see Section 2.3.2). The N 

in cereal crops is typically derived from added fertilizer 

(synthetic or manures) and from mineralization of soil 

organic matter (conversion of complex organic molecules 

to ammonium) in about equal amounts. As discussed 

in Section 2.2 and Section 5.3.4, crop production is not 

efficient in using Nr so only 30-70% (a global average 

of 40%) of all the N mobilized for crop production is 

harvested in the crop. The remainder is in crop residue 

(roots and above ground stover) stored in the soil, leached 

to aquatic systems as NO3

-

, volatilized to the atmosphere 

as NH3 or NOx or denitrified (see Section 4.8, Figure 18) 

to produce NOx, N2O and N2. An additional ~1.1 Tg of 

synthetic fertilizer N is used to maintain turfgrass in the 

urban environment (see Section 2.2.5) and another 0.1-0.2 

Tg N is used to enhance forest production.

Within the nitrogen cascade (Figure 1), the interactions 

between the agricultural and populated portions of the 

terrestrial system dictate the production and flow of Nr. 

Although occupying the largest area, forest and grassland 

portions of terrestrial ecosystems serve mainly to absorb 

atmospheric deposition and provide a source of forest 

products and forage for livestock production. Reactive 

nitrogen input into these systems is from biological N 

fixation in unmanaged lands, atmospheric deposition, and 

Nr from livestock manure that is deposited. The livestock 

that is grazing within grasslands (Table B-1) may lead 

to the N saturation of unmanaged forest and grassland 

ecosystems (Galloway et al., 2004; Bobbink et al., 2010). 

This report uses the Nr input numbers from Table B-1 

and food production numbers to estimate the flow of Nr 

through agricultural and populated parts of the terrestrial 

system (Table B-2). The FAO (20010b); www.fao.org/

es/ess/top/country.html ) lists the 20 largest agricultural 

commodities produced, imported, and exported in the 

U.S. in 2002. Of these commodities, corn (229 Tg), 

soybeans (75 Tg), wheat (44 Tg), and cow’s milk (77 Tg) 

were produced in the greatest amount. Using commodity 

N content data derived from data used to calculate crop 

Table B-1: Sources of reactive N into terrestrial systems in the United States in 2002 (from Table 1 

data sources; in Tg N/yr).

Source Environmental System

Agricultural Vegetated Populated Total

Forest Grassland

Atmospheric 1.3 1.4 1.9 0.4 6.9*

N fixation 7.7 – 6.4 – 14.1

Synthetic N 9.7 0.1 ** 1.1 10.9

Animal manure## 1.2 – 3.8# – 6.0#

Human sewage## 0.1 – – 1.2 1.3

*The amount of atmospheric Nr deposition is based on area of each environmental system within the continental U.S. 

The total area does not sum to 100% because non-arable lands are not included in this table. 

**Synthetic fertilizer N used for managed pasture fertilization is included in the agricultural land classification.

#Unrecoverable livestock manure deposited on grasslands, the unaccounted for ~1 Tg of Nr assumed to be lost 

through ammonia volatilization, leaching, or denitrification (U.S. EPA, 2007e).

##Note that livestock manure and human sewage used as fertilizer are recycled N components of the nitrogen cascade 

and not new Nr inputs.
96

residue N in the EPA inventory of U.S. greenhouse gas 

emissions and sinks (U.S. EPA, 2007e), an estimated 9 

Tg of N was marketed in three crops, soybeans (4.4 Tg 

N; from U.S. EPA, 2007g), corn (3.2 Tg N), and wheat 

(0.9 Tg N). Whole milk contained ~ 0.5 Tg of N while 

other meat and egg produce contained ~1.4 Tg of N, 

totaling ~ 1.9 Tg N. Grain, fruits, nuts and vegetables 

contained ~9.3 Tg of N. If the total N input use efficiency 

is 40%, then ~23 Tg of N from all sources is required to 

produce 9.3 Tg of vegetative commodities. Table B-2 

lists the estimated Nr input into agricultural systems (~ 

20 Tg) and additional N input from crop residue that was 

returned to the field the previous year (4.4 Tg) and from 

mineralization of soil organic matter (4.7 Tg). All of this 

N input totals ~29 Tg of N that is actually involved in 

the production of the 9.3 Tg of crop commodity N. If one 

assumes that return of crop residue to the field is directly 

proportional to crop production, then 24.3 Tg of N was 

required to produce the 9.3 Tg of crop commodity N. 

These estimates indicate that ~38% of the total annual 

input of N that went into the agricultural crop production 

system was contained in the main crop commodities 

produced in the U.S. in 2002.

Of the 24.3 Tg N required to produce crop commodity 

N in 2002, approximately 2.5 Tg was used to grow feed 

used for milk, egg, and meat production. This estimate is 

made assuming that 4 units of N are required to produce 

a unit of milk, eggs, or meat (see Section 5.3.4). This 

estimate also assumes that one-third of N required for 

livestock production comes from commodities in the 

FAO top-20 list and the remaining two-thirds comes 

from alfalfa, silage, and grass over the course of a year 

(Oitjen and Beckett, 1996). Approximately 4.3 Tg of N in 

agricultural commodities (2.8 Tg in soybeans, corn, and 

wheat) were exported, while ~0.15 Tg N was imported 

in various food and drink commodities. The U.S. human 

populace consumed ~1.96 Tg of N in 2002 (292 million 

people, consuming 114.7 g protein/person/day, 0.16 g N/g 

protein, 365 days) (approximately 1.2 Tg from animal 

protein-N and 0.7 from vegetative protein).

These three consumption areas – internal consumption 

of vegetable N for livestock production, human 

consumption, and export – account for 77% of the 

commodities produced. The unaccounted for commodity 

N is likely partly in annual storage. Some smaller fraction 

of annual production is used for pet food and a small 

fraction is returned to the terrestrial environment because 

of spoilage and handling losses.

In forests and grasslands (vegetated system) N input 

in 2002 was ~3.5 Tg of anthropogenically introduced 

N, with the remaining ~10.1 Tg derived from BNF and 

livestock manure deposition. Of this anthropogenic N, 

~21% was retained in soil and tree biomass, while the 

remainder was removed in tree harvest (~0.2 Tg, see 

Section 2.3.2.) or lost to other parts of the environment 

through NH3 volatilization and NO3

– leaching and runoff 

(Table B-2). Total N input into agricultural systems 

was ~20 Tg, with ~ 11 Tg being removed as products 

including the transfer of ~2 Tg N as food to the human 

population. Almost 40% of the N input into agricultural 

systems is lost through NH3 volatilization, nitrification/

denitrification, and NO3

-

 runoff. The 4.2 Tg of Nr of 

Haber-Bosch N that is used for industrial feedstock is 

not included in this assessment. Of the input of ~3.3 Tg 

of N into the populated system ~80% is lost through 

human excreta processed in sewage treatment plants, 

denitrification in soils, and leaching and runoff of NO3

-

(Table B-2).

Table B-2 summarizes the input and flow of Nr in 

the main terrestrial systems within the continental U.S. 

Anthropogenic input of Nr into forests and grasslands 

totaled ~3.5 Tg in 2002, with an estimated 6.4 Tg of Nr 

being introduced through natural biological N fixation. 

Of this Nr ~ 0.7 Tg was stored in vegetation and soils 

(see Section 2.3.2) and ~2 Tg removed as livestock 

forage, while the remainder was lost to the atmosphere 

and aquatic systems, or removed as forest products and 

livestock forage. The largest anthropogenic Nr input (~20 

Tg) was into agricultural production where ~11.2 Tg was 

removed as agricultural product, ~ 2 Tg transferred as 

Table B-2: Nr input and flows (Tg N/yr) in the terrestrial portion of the Nitrogen Cascade (Figure 1) 

within the continental United States in 2002

Environmental 

System*

N Input to 

System

N Storage in 

System**

Agricultural & 

Forest 

Products

Transfers to 

Aquatic or 

Atmospheric

Transfers as 

a % of Input

Vegetated 13.6 0.7 2.2 10.7 79

Agricultural 19.6 0.8 11.2 7.6 39

Populated 3.3 0.1 0 3.2 97

*The Environmental Systems are those noted in the terrestrial portion of the N cascade shown in Figure 1. Data from 

Table B-1, derived from regrouping information from Table 1 data sources, are shown in Table B-2.

**Estimates are from section 2.3.2 of this report.
97

edible product to the “populated” portion of the terrestrial 

system, ~0.8 Tg was stored in agricultural lands, and ~7.6 

Tg N was lost to the atmosphere and aquatic systems. 

New N input into the “populated” portion totaled ~3.3 Tg, 

which came from N transfer in food and use of fertilizer 

N in lawns, gardens and recreational areas. Within these 

areas an estimated 0.12 Tg was stored in urban forests. 
98
99

For various reasons, wetland restoration has been 

proposed and the magnitude of needed restoration 

estimated. For the Wetland Reserve Program (WRP), the 

Farm Bill of 1990 set a goal of restoring approximately 

1 million acres. A few years later, the NRC (1992) 

proposed a national goal of restoring 10 million acres of 

inland and coastal wetlands by 2010. The NRC went on 

to recommend that 400,000 miles of streams and rivers 

be restored by 2012 and that 1 million acres of lakes 

be restored by 2000, both of which would further the 

control of reactive nitrogen. While none of these goals 

has been or is likely to be met by the recommended 

date, they articulated a need for wetland restoration 

addressing the important relationship between wetlands 

and water quality. 

Taking into account the economics of using wetlands 

to manage Nr adds yet another dimension to site 

selection. Based on the results of the Water Environment 

Research Foundation’s study (Hey et al., 2005a,b), the 

Kinship Foundation sponsored a study (Scott et al., in 

preparation) to define the market for producing and 

selling Nr (as NO3

–) credits. For this analysis, a real 

potential market area was selected: the Illinois River 

watershed in Illinois – the tributaries draining Wisconsin, 

Indiana, and Michigan were excluded. The producers 

of nitrogen credits were identified as “nutrient farmers” 

and they became the “sellers” of N credits. The “buyers” 

of nitrogen credits were restricted to municipal and 

industrial wastewater treatment facilities, those facilities 

that hold an NPDES permit. This restriction, of course, 

resulted in a considerable understatement of the market 

size since the identified buyers emit less than 11% of the 

total aquatic N load (David and Gentry, 2000) that finds 

its way to the Mississippi River – air emission/deposition 

and agriculture account for the remaining 89%. 

The watershed was divided into 19 sub-watersheds, 

spatially locating credit supply and demand. A linear 

programming model was developed and used to: (1) 

examine the potential extent and distribution of nitrogen 

credit demand and supply; (2) compare the average 

seasonal demand levels to the supply capacity of nutrient 

farms; and (3) evaluate the relative effects of seasonality. 

Market efficiency was imposed through the objective 

function: the least costly distribution of credit production 

to meet the given monthly demand. Thereby, sellers 

and buyers were identified and linked and the spatial 

characteristics of the market mapped by sub-watershed. 

At the same time, the equilibrium price of a credit, or the 

prevailing price at which buyers and sellers are willing 

to trade, was determined. The market, as represented by 

the model, determined where the most intensive wetland 

investment (i.e., wetland restoration) would be, the 

revenues returned to these investments, and the costs and 

savings to the buyers.

All 290 permitted buyers are geographically 

distributed as shown in Figure C-1. The mass loading 

of the buyers (2,423 tons/month) is reflected in Figure 

C-2. Eighty-nine percent of the demand comes from the 

northeastern corner of the Basin (Upper Fox, Des Plaines, 

and Chicago/Calumet sub-watersheds), the Chicago 

metropolitan area. As illustrated by Figure C-3, 41% of 

the wetland restoration area (using the criteria discussed 

above) was identified in the southwestern corner of the 

watershed (Lower Illinois, La Moine, Macoupin, Lower 

Sangamon, and Middle Illinois sub-watersheds), where 

the floodplain is almost entirely leveed. For the market 

study, the available load of Nr (NO3

-

) by season and 

sub-watershed was mapped as illustrated in Figure C-4. 

The N load was computed using water quality and flow 

data collected by the U.S. Geological Survey from 1987 

to 1997. The wetland and wastewater cost functions are 

described in Hey et al., 2005; however, the wetland cost 

functions were modified for the market study to reflect the 

variability of land costs across the watershed (i.e., higher 

land values in urban Chicago and lower land values in 

rural Illinois). This variability is reflected in the spatial 

distribution marginal costs shown for the spring marginal 

costs depicted in Figure C-5. Wetland treatment costs vary 

by time of year because the level of microbial activity, 

which drives the denitrification process, varies with water 

temperature. Therefore, treating an equivalent load of Nr 

requires more wetland area in winter than in summer. 

Three regulatory scenarios

Regulatory agencies may require that dischargers and 

nutrient farms be located in proximity to each other and 

could impose “penalties” when the two are not. Thus, 

for the sake of analysis, the Committee created three 

regulatory scenarios: 

1. Unrestricted – buyers can purchase nitrogen credits 

from nutrient farmers anywhere in the watershed 

without regard to location. The result of this scenario is 

given in Figure C-6.

2. Restricted intra-watershed – the buyer must purchase 

all available credits within its own sub-watershed 

before buying in other sub-watersheds

Appendix C:

Water Quality Trading in the Illinois 

River Basin
100

3. Accrued 10% penalty – buyers pay an increasing 

“tax” on credits purchased in consecutive downstream 

watersheds 

The three regulatory scenarios were analyzed for each 

of the four seasons. (D. Hey, Wetlands Research, Inc., 

Personal Communication.)

The “unrestricted” scenario is the least expensive 

because nutrient farms in this scenario are located 

downstate, where land is least expensive. In the other two 

scenarios, credits were purchased a little more evenly 

throughout the watershed. Still, most of the credits in the 

southern corner of the watershed were purchased. The 

Figure C-1: Distribution of municipal discharge 

and industrial dischargers in the Illinois River 

Watershed

Municipal dischargers shown are those exceeding one 

million gallons per day. Symbols may represent more 

than one discharger at that location.

Source: D. Hey, Wetlands Research, Inc., Personal 

Communication

Figure C-2: Distribution of total nitrogen 

emissions by sub-watershed

Source: D. Hey, Wetlands Research, Inc., Personal 

Communication

Figure C-3: Potential land availability in the 

100-year flood zone for nutrient farming in each 

sub-watershed in the Illinois River watershed

Source: D. Hey, Wetlands Research, Inc., Personal 

Communication

Figure C-4: Spring available total nitrogen load 

by sub-watershed

Source: D. Hey, Wetlands Research, Inc., Personal 

Communication
101

“restricted intra-watershed” and “accrued 10% penalty” 

scenarios resulted in more credits being purchased. This 

resulted in the sale of N credits exceeding the mass 

of Nr emitted by wastewater treatment, which would 

benefit the overall control of reactive nitrogen. It also 

would increase the value of the market and the profits 

of the nutrient farmer. The downside of such regulatory 

controls is that they would drive up the effective price of 

nitrogen credits. If a buyer had to buy a 1.5 tons for every 

ton discharged because credits are not available in the 

tributary watershed, the effective price of a credit would 

be 1.5 times the price of the tributary sub-watershed. If 

prices rise too much, “concrete and steel” technologies 

may become competitive.

Considering all of the point source dischargers 

in the Illinois River watershed, between 29,000 and 

36,000 tons TN/year could be removed through nutrient 

farming under the studied trading schemes (Table C-1). 

The range of removal is a function of the penalties 

imposed on the market by the regulatory agencies. 

Accordingly, the market revenue would range from 

$70 million to $121 million/year. This is a sizeable 

market that could generate substantial profits, from $6 

million to $38 million with the return on investment 

varying from 5 to 25%. If the savings are shared evenly 

between the seller and buyer, the nutrient farmer could 

earn between $200 and $300/acre/year net profit, which 

in many cases is greater than the profits from corn 

or soybean production. Further, these profits do not 

include any earnings from flood control or recreation, as 

suggested in a McKnight study report (Hey et al., 2004). 

With such profits, sufficient land should be available for 

nutrient farming.

This analysis indicates that appropriate lands 

are available and that wetlands can be effectively 

restored and efficiently used to control Nr. The market, 

structured as discussed above, could generate the capital 

to accomplish the needed large-scale wetland restoration 

while saving taxpayers the cost of upgrading their 

municipal wastewater treatment plant (TWI, 2007).

Source: D. Hey, Wetlands Research, Inc., Personal 

Communication

Figure C-5: Spring marginal cost (price) by 

watershed Figure C-6: Unrestricted spring credit sales 

(tons/month) by sub-watershed Source: D. Hey, Wetlands Research, Inc., Personal 

Communication

Table C-1: Nutrient farm market parameters under three trading scenarios

Source: D. Hey, Wetlands Research, Inc., Personal Communication

Parameter Unrestricted Restricted 

Intra-watershed

Accrued 10% 

Penalty

Total Credits Sold (tons) 29,078 29,078 35,781

Total Revenue 21 $69,925,497 $99,571,889 $121,457,652

Total Cost to Produce Credits $63,258,006 $66,193,924 $83,288,747

Profit $6,667,491 $33,377,968 $38,168,905

21 Assumes all credits were sold at the cheapest cost within the Illinois River watershed.
102
103

The European Union has undertaken broad measures, 

based on the critical-loads concept, to manage Nr. 

Tables D-1, D-2, and D-3 summarize several different 

environmental impacts currently used as indicators 

and identify whether there are current limit values 

set by the United Nations Economic Commission for 

Europe (UNECE) or European Union (EU). These 

tables identify the main links to the cascade of reactive 

nitrogen in the environment, the relevance and link to 

Nr of the effect/pollutant, and existing agreements in 

which the effect is currently addressed. In addition, 

some impacts are more relevant than others in relation 

to societal importance and the connection to the nitrogen 

cascade. The categorization on a scale of 1 (highest 

relevance) to 5 (unimportant) provides a first-level 

prioritization for future mitigation activity. The last 

column summarizes existing links to international 

regulations and conventions. 

Where there is a limit and the relevance for the 

nitrogen cascade is high, then this might be the limiting 

factor for Nr production and its associated losses to the 

environment. Some limits might be more relevant in 

specific areas and less relevant in others. For example 

NO2 concentrations relevant for human health are 

limited to 40 ppb in urban areas, limiting industry and 

traffic, but would probably not be an issue of concern 

in remote areas with low population densities. In 

these areas, however, loss of biodiversity might limit 

nitrogen deposition and therewith the sources in the 

region. The only way to determine the extent that 

critical thresholds are limiting is by overlaying them 

on different regions and determining, through the use 

of monitoring data or by modeling exercises, where 

and which sources contribute to exceeding the critical 

threshold. Then the best methods for putting caps on 

relevant sources can be identified. A pre-classification 

of regions might be useful, e.g., urban regions, remote 

regions, marine areas, etc. One aspect of this global 

view of nitrogen impacts and metrics that is evident is 

the mix of “classical-” and “service”-based categories, 

consistent with the need for an integrated approach to 

the management of nitrogen. 

Appendix D:

Management of Nr Measures Based on the 

Concept of Critical Loads
104

Metrics Regulated? Link to Nr 

cascade Relevance*

Regulatory or 

political 

convention

Respiratory disease in people 

caused by exposure to high 

concentrations of:

Ozone Sum of ozone 

over 35 ppb 

YES NOx

emissions

3 Convention on Longrange Transboundary Air 

Pollution 

Clean Air for Europe

Other photochemical 

oxidants

Org. NO3, PAN 

concentration 

(atm)

NO NOx

emissions

5 Convention on 

Long-range 

Transboundary Air 

Pollution et al.

Fine particulate 

aerosol

PM10, PM2.5 

concentration 

(atm)

YES NOx, NH3 em 1 Convention on Longrange Transboundary Air 

Pollution

Clean Air for Europe

Direct toxicity of nitrite 

NO2

-

NO2

-

concentration

YES NOx 2 World Health 

Organization 

Convention on Longrange Transboundary Air 

Pollution

Clean Air for Europe

Nitrate contamination of 

drinking water

NO3

-

concentration 

(aq.)

YES NO3

-

leaching

2 EU Essential 

Facilities Doctrine 

Depletion of stratospheric ozone NOx, N2O 

concentration/

flux (atm)

NO NOx, N2O 3 Montreal Protocol

Increase allergenic pollen 

production, and several 

parasitic and infectious human

NO 5 None

Blooms of toxic algae and 

decreased swimability of 

in-shore water bodies

Chlorophyll a

NO3

-

 (&P) 

concentration 

(aq)

NO Runoff, Nr 

deposition

1 Convention for the 

Protection of the Marine 

Environment of the 

North-East Atlantic

Helsinki Commission

Barcelona Convention

Table D-1: Summary of the effects of excess Nr on human health in relation to metrics, current 

international regulations and conventions, and the link to the nitrogen cascade

atm – atmospheric; aq – aqueous

*Relevance and link to nitrogen incorporates societal priority and N contribution: 1. highest relevance, 2. high 

relevance, 3. significant relevance, 4. some relevance, 5. unimportant. 
105

Table D-2: Summary of the effects of excess Nr on ecosystems related to currently used metrics, the 

existence of European regulatory values, and the link to the nitrogen cascade.

Metrics Regulated? Link to Nr 

cascade Relevance*

Regulatory or 

political 

convention

Ozone damage to crops, 

forests, and natural 

ecosystems

AFstY (O3 flux), 

AOT40** 

YES NOx 2 Convention on Longrange Transboundary 

Air Pollution 

Clean Air for Europe

Acidification effects on 

terrestrial ecosystems, ground 

waters, and aquatic 

ecosystems

Critical loads YES Nr deposition 2 Convention on Longrange Transboundary 

Air Pollution 

Clean Air for Europe 

WFD

Eutrophication of 

freshwaters, lakes 

(incl. biodiversity)

Biological 

Oxygen Demand, 

NO3

-

 conc (aq)

Critical loads

YES

NO

Runoff, Nr 

deposition

3 Water Framework 

Directive

Eutrophication of coastal 

ecosystems inducing 

hypoxia (incl. biodiversity)

BOD, NO3

-

 conc 

(aq)

Critical loads

BOD, NO3

-

 

conc (aq)

Critical load

Runoff, Nr 

deposition

1 Convention for the 

Protection of the 

Marine Environment of 

the North-East Atlantic

Helsinki 

Commission

Barcelona 

Convention

Nitrogen saturation of soils 

(incl. effects on GHG 

balance)

Critical loads YES Nr deposition 1 Convention on Longrange Transboundary 

Air Pollution 

Clean Air for Europe

Biodiversity impacts on 

terrestrial ecosystems 

(incl. pests and diseases)

Critical loads, 

critical levels 

(NH3, NOx)

YES Nr deposition 1 Convention on Longrange Transboundary 

Air Pollution 

Clean Air for Europe

Convention on 

Biological Diversity

atm – atmospheric; aq – aqueous

*Relevance and link to nitrogen incorporates societal priority and N contribution: 1. highest relevance, 2. high 

relevance, 3. significant relevance, 4. some relevance, 5. unimportant. 

**Accumulated ozone exposure over a threshold of 40 parts per billion
106

Table D-3: Summary of the effects of excess N on other societal values in relation to metrics 

and regulatory values in current international regulations and conventions and the link to the 

nitrogen cascade

Metrics Regulated? Link to Nr 

cascade Relevance*

Regulatory or 

political 

convention

Odor problems associated with animal 

agriculture

Acidity in precipitation., 

O3,

PM

YES NOx, NH3 3 Convention on 

Long-range 

Transboundary Air 

Pollution 

Effects on 

monuments and 

engineering 

materials

PM2.5 conc (atm) NO NOx, NH3 4

Global climate 

warming induced by 

excess nitrogen

N2O, conc/flux (atm) NO NOx, NH3 1 United Nations 

Framework 

Convention on Climate 

Change 

Regional climate 

cooling induced by 

aerosol)

PM2.5 conc (atm) NO NOx, NH3 1 United Nations 

Framework 

Convention on Climate 

Change

atm – atmospheric; aq – aqueous

*Relevance and link to nitrogen incorporates societal priority and N contribution: 1. highest relevance, 2. high 

relevance, 3. significant relevance, 4. some relevance, 5. unimportant. 
107

Production of N2 and N2O via gas-phase 

reactions 

Atmospheric conversion of NOx and NHx to less 

reactive N2 or N2O appears to play a minor role in the 

global N budget, but currently is not well quantified. The 

gas-phase reactions in the troposphere that convert NH3

and NOx to N2 and N2O, start with attack of NH3 by OH: 

NH3 + OH· → NH2· + H2O (1)

Several potentially interesting fates await the NH2

radical:

NH2· + O3 → NH, NHO, NO (2)

NH2· + NO2 → N2O + H2O (3)

NH2· + NO → N2 + H2O (4)

kO3 = 1.9x10-13 cm3 s-1

kNO2 = 1.8x10-11 cm3 s-1

kNO = 1.8x10-11 cm3 s-1

The first step, attack by OH, is slow. The rate constant 

for the Reaction 1 is 1.6 x 10-13 cm3 s-1 and the lifetime 

of NH3 for a typical concentration of 106 OH cm-3

is about 70 days. In most areas of the world where 

concentrations of NH3 are high, concentrations of sulfates 

are also high, and NH3 is removed by conversion to 

condensed phase ammonium sulfate or bisulfate on time 

scales much faster than 70 d. The mean lifetime of these 

aerosols with respect to wet deposition is about 10 d.

There are some areas of the world, notably California 

and South Asia, where NH3 and NOx are emitted in large 

quantities, but SO2 is not, and there gas-phase conversion 

can take place. NH3 is usually removed by wet or dry 

deposition. Reaction 2 is relatively slow and oxidation of 

ammonia is in general an unimportant source of NOx. But 

Reactions 3 and 4 may be atmospherically noteworthy. As 

an upper limit to current N2O production, we can assume 

that each of these regions covers an area of 106 km2 and 

that they contain ammonia at a concentration of 10 g N m-3

in a layer 1,000 m deep. The annual production of N2 and/

or N2O would then be on the order of 0.1 Tg N, a minor 

but nontrivial contribution to denitrification and about 1% 

of the anthropogenic N2O production. If NH3 rich air is 

lofted out of the boundary layer into the upper troposphere 

where deposition is impeded, it will have an atmospheric 

residence time on the order of months, and the probability 

of reaction to form N2O or N2 becomes greater. This 

possibility has not been investigated extensively. It is also 

possible than Europe and North America will continue to 

reduce S emissions without reducing NH3 emissions and 

the atmospheric source of N2O will grow in importance.

In the stratosphere, N2O photolysis leads to loss of 

Nr via

N2O + hμ → N2 + O (5)

While reaction with an electronically excited oxygen 

atom O(1D) leads to production of NO via

N2O + O(1D) → 2NO (6)

Photolysis (Reaction 5) dominates, but a large 

enough fraction of the N2O reacts with O(1D) that this 

is the main source of NOx in the stratosphere. The fate 

of this oxidized nitrogen (NOy) is transport back into 

the troposphere where it is removed by wet deposition. 

Downward transport of the odd N from the oxidation of 

N2O is a minor (~1%) source of NOy in the troposphere. 

Most of the N2O released into the atmosphere is 

eventually converted to N2 – the problem is that it 

destroys stratospheric ozone in the process. 

In summary, our current understanding of the 

chemistry of atmospheric ammonia suggests that in situ 

conversion to N2 and N2O plays a minor (~1%) role 

in global N budgets, but if assumptions about kinetics 

or concentrations are in error, these mechanisms could 

become important.

SPARROW model for estimating watershed Nr

Estimates of Nr transfers in aquatic ecosystems are 

difficult to quantify at the national scale, given the need 

to extrapolate information from sparse monitoring data 

in specific watersheds to the geographic boundaries of 

the nation. One excellent tool for estimating Nr loads at 

regional scales is the spatially referenced regression on 

watershed attributes (SPARROW) modeling technique. 

The SPARROW model has been employed to quantify 

nutrient delivery from point and diffuse sources to 

streams, lakes, and watershed outlets at the national scale 

(Smith et al., 1997). The model infrastructure operates 

in a geographic framework, making use of spatial data 

to describe sources of pollutants (e.g., atmospheric 

deposition, croplands, fertilizers) and characteristics of 

the landscape that affect pollutant transport (e.g., climate, 

topography, vegetation, soils, geology, and water routing). 

Though empirical in nature, the SPARROW modeling 

approach uses mechanistic formulations (e.g., surfacewater flow paths, first-order loss functions), imposes mass 

balance constraints, and provides a formal parameter 

estimation structure to statistically estimate sources and 

fate of nutrients in terrestrial and aquatic ecosystems. The 

spatial referencing of stream monitoring stations, nutrient 

sources, and the climatic and hydrogeologic properties 

of watersheds to stream networks explicitly separates 

Appendix E:

Technical Annexes
108

landscape and surface-water features in the model. This 

allows nutrient supply and attenuation to be tracked 

during water transport through streams and reservoirs, 

and accounts for nonlinear interactions between nutrient 

sources and watershed properties during transport. The 

model structure and supporting equations are described 

in detail elsewhere (Smith et al., 1997, Alexander et al., 

2000, Alexander et al., 2008). Figure E-1 provides an 

estimate of contemporary Nr loading in surface waters 

of the U.S., representing long-term average hydrological 

conditions (over the past three decades). There are hot 

spots of high Nr yields to rivers associated with land use 

and watershed characteristics, and SPARROW allows 

considerations of the fate of these Nr inputs to streams 

and rivers as they flow downstream to coastal receiving 

waters (Alexander et al., 2008).

Figure E-1: Total Nr yields (kg/ha/yr) in large rivers of the U.S. 

Data Source: Alexander et al., 2008
109

EPA informed the Committee that it is in the process 

of implementing a number of regulations to reduce NOx

from a variety of mobile sources22. These include clean 

diesel regulations for trucks and buses and nonroad 

engines, as well as locomotives and smaller marine 

vessels. EPA first regulated NOx emissions from motor 

vehicles for the 1973 model year and since then has 

tightened these standards. EPA’s efforts to control NOx

emissions from nonroad vehicles, locomotives, and 

commercial marine vessels started in the 1990s. NOx

reductions for each rule were calculated by EPA based 

on inventories available at the times of the rules.

1. Light Duty Tier 2 Rule – EPA’s Tier 2 Vehicle and 

Gasoline Sulfur Program (65 FR 6698, February 10, 

2000). This program requires new cars, sport utility 

vehicles (SUVs), pickup trucks, and vans to be 77 

to 97% cleaner than 2003 models, while reducing 

sulfur levels in gasoline by 90%. EPA estimates that 

as newer, cleaner cars enter the national fleet, the new 

tailpipe standards will reduce emissions of nitrogen 

oxides from vehicles by 3 million tons, or about 

74% in 2030. Prior to that, the EPA Tier 1 vehicle 

regulations, effective with the 1995 model year, also 

resulted in significant NOx reductions.

2. EPA’s Clean Heavy Duty Truck and Bus Rule. When 

the Agency finalized the Heavy Duty Truck and Bus 

Diesel Rule (66 FR 5002, January 18, 2001) in 2001, 

trucks and buses accounted for about one-third of 

NOx emissions from mobile sources. In some urban 

areas, the contribution was even greater. With model 

year 2010, all new heavy duty trucks and buses will 

result in NOx emission levels that are 95% below 

the pre-rule levels. EPA projects a 2.6 million ton 

reduction of NOx emissions in 2030 when the current 

heavy-duty vehicle fleet is completely replaced with 

newer heavy-duty vehicles that comply with these 

emission standards.

3. Clean Air Nonroad Diesel – Tier 4 Rule (69 FR 

38957, June 29, 2004). In 2004, EPA adopted a 

comprehensive national program to reduce emissions 

from future nonroad diesel engines by integrating 

engine and fuel controls as a system to gain the 

greatest emission reductions. EPA estimates that in 

2030, this program will reduce annual emissions of 

NOx by about 740,000 tons.

4. Marine-Related NOx Reductions from 1999 to 2003. 

EPA completed three rulemakings with respect to the 

diesel marine sector that will reduce NOx emissions. 

These rules are now in effect and being phased-in. 

In 1999 (64 FR 73299, December 29, 1999), EPA 

promulgated NOx requirements for diesel engines 

used in commercial boats (large inland and nearshore boats) and commercial vessels (ocean-going 

vessels). EPA estimates that these reduced emissions 

from these vessels by about 30%. In 2002 (67 FR 

68241, November 8, 2002), EPA promulgated rules 

reducing NOx emissions from diesel engines used in 

recreational marine vessels by 25%. In 2003 (68 FR 

9746, February 28, 2003), EPA promulgated another 

rule further reducing NOx from diesel engines used in 

commercial vessels by about 20%. EPA projects that 

on a nationwide basis, these four programs will reduce 

marine-related NOx by more than 1 million tons in 

2030.

5. Locomotive and Marine Diesel Rule (73 FR 25098, 

May 6, 2008). In March 2008, EPA adopted standards 

that will reduce NOx emissions from locomotives 

and marine diesel engines. The near-term emission 

standards for newly-built engines phased in starting 

in 2009. The long-term standards begin to take effect 

in 2015 for locomotives and in 2014 for marine diesel 

engines. EPA estimates NOx emissions reductions 

of 80% from engines meeting these standards. EPA 

projects that in 2030, about 420,000 tons of NOx will 

be reduced from the locomotive engines, and 375,000 

tons of NOx will be reduced from commercial and 

recreational marine engines.

6. Non-road Spark-Ignition Engines (73 FR 59034, 

October 8, 2008). In 2002, EPA promulgated emissions 

standards for large spark-ignition engines. These took 

effect in 2004 for Tier 1 standards and in 2007 for Tier 

2 standards. EPA promulgated emissions standards 

for small spark-ignition engines in 2008. EPA projects 

that, when fully implemented, the new standards will 

result in a 35% reduction in HC+NOx emissions from 

new engines’ exhaust, reduce evaporative emissions by 

45%, and that together these programs will reduce NOx 

by more than 585,000 tons in 2030.

7. EPA’s Coordinated Strategy for Control of Emissions 

from Ocean-Going Vessels (www.epa.gov/otaq/

oceanvessels.htm). EPA’s coordinated strategy to 

Appendix F:

Recent Major EPA Mobile Source Rules to 

Control NOx

22 The information in Appendix F was provided to the Integrated Nitrogen Committee by Mazrgaret Zawacki of the U.S. EPA Office of Transportation 

and Air Quality.
110

control emissions from ocean-going vessels consists 

of actions at the national and international levels. 

On December 22, 2009, EPA finalized emissions 

standards for ocean-going vessels which take effect 

in 2011. In addition to this rule the U.S. Government 

has also amended MARPOL Annex VI to designate 

U.S. coasts as an Emission Control Area (ECA) in 

which all vessels, regardless of flag, will be required 

to meet the most stringent engine and marine fuel 

sulfur requirements in Annex VI. New engine emission 

and fuel sulfur limits contained in the amendments to 

Annex VI are also applicable to all vessels regardless 

of flag and are implemented in the U.S. through the Act 

to Prevent Pollution from Ships (APPS). EPA projects 

that when fully implemented, the coordinated strategy 

will reduce NOx emissions from ocean-going vessels 

by 80% and that in 2030, the coordinated strategy 

is expected to yield a reduction in NOx of about 1.2 

million tons.

8. EPA’s Voluntary Clean Diesel Programs. EPA has 

created a number of programs designed to reduce 

emissions (including both PM and NOx) from the diesel 

fleet. In conjunction with state and local governments, 

public interest groups, and industry partners, EPA has 

established a goal of reducing emissions from the over 

11 million diesel engines in the existing fleet by 2014. 

Looking at these engines, EPA determined there were 

general sectors that provided the best opportunity to 

obtain significant reductions and created programs for 

Clean Agriculture, Clean Construction, Clean Ports, 

Clean School Bus, and SmartWay Transport.

9. Section 177 of the Clean Air Act allows states outside 

of California to adopt California emissions standards, 

once EPA has granted such a waiver. As a result, 

several northeastern states have adopted California 

standards. Maryland adopted its California LEV II 

NOx standards as part of its Low Emission Vehicle 

Program (COMAR 26.11.34, effective December 17, 

2007). These standards take effect with the 2011 model 

year. Maryland submitted that program to EPA as a 

SIP revision. Pennsylvania adopted California LEV II 

NOx standards as part of its Clean Vehicles Program 

(codified at Pa. Code Chapters 121 and 126, effective 

December 9, 2006). Pennsylvania’s program began 

with model year 2008 vehicles. Pennsylvania submitted 

this program as a SIP revision.
111

Nitrogen contamination of groundwater

In addition to environmental concerns about 

N-nutrient loading to freshwaters from a groundwater 

pathway, there are also potential human health impacts 

from elevated levels of N in groundwater, especially 

from NO3. It has been long established that excess NO3

in drinking water supplies can cause blue baby syndrome 

(methemoglobinemia) (Knobeloch and Proctor 2001; 

Ward et al. 2005, 2006), the indicator of which is MetHb. 

To protect public health from effects of NO3 in drinking 

water, EPA has established a maximum contaminant level 

goal (MCLG) that considers a lifetime exposure plus a 

margin of safety. For NO3 in drinking water, the MCLG 

is 10 mg/L. Nitrite-N also has an established MCLG of 

1 mg/L, and the combined NO3 and NO2 MCLG set by 

EPA is 10 mg/L. These same values are used to regulate 

NO3 and NO2 as maximum contaminant levels (MCL), 

which are the highest levels of contaminants allowed in 

drinking water (40 CFR § 141.62). 

The drinking water standard is commonly exceeded 

in streams and rivers of the U.S., particularly in the 

agricultural Midwestern U.S. For example, there were 13 

episodes over a 25- year period of formal warnings by 

authorities to local citizens in Columbus, Ohio, about not 

drinking tap water because nitrate-nitrogen was higher 

than 10 mg-N/L (Mitsch et al., 2008). These episodes 

lasted from one to several weeks each. The pattern is 

generally for high concentrations of nitrate-nitrogen 

in Midwestern rivers from February through June or 

July. In one pattern, averaged over 7 years with weekly 

river sampling, nitrate-nitrogen in a central Ohio river 

peaked with an average of 7 mg-N/L in June after which 

concentrations decrease to 1-2 mg-N/L for the rest of the 

summer and fall. The year-to-year variability was high 

for that month as well (Mitsch et al. 2005) as these spring 

“high-nitrate” floods do not occur every year. The nitrate 

“pulses” generally are part of flood events after fertilizer 

has been applied to fields in the watershed. 

Public policy by water supply agencies is to treat 

high concentrations of nitrate-nitrogen in drinking water 

supplies as a real public health threat. Recent studies 

have brought the concern of high nitrate-nitrogen in 

drinking water into dispute (Ward et al., 2005, 2006). 

While there is a definite link between excessive nitrate 

in drinking water and methemoglobinemia, there is also 

a need to better understand the interaction of the range 

of environmental factors (e.g., cofactors such as diarrhea 

and respiratory diseases reportedly increase MetHb 

levels) that promote methemoglobinemia. This will 

help identify the environmental conditions under which 

exposure to nitrate in drinking water poses a risk of 

methemoglobinemia.

According to the USGS (Barber, 2009) Summary of 

Estimated Water Use in the U.S. in 2005, total water 

withdrawals in the U.S., excluding thermoelectric power 

usage, were 210 billion gallons per day, of which 44,200 

million gallons per day (MGD) were for public water 

supply. About two-thirds of that supply is provided by 

surface water, the rest is from wells and about 58% 

(25,600 MGD) of public water supply goes towards 

domestic use, including drinking water. Private wells 

(Figure G-1) that are not part of public water supply 

systems are estimated to provide an additional 3,830 

MGD, providing domestic water for 42.9 million people 

(14% of the U.S. population in 2005). 

Groundwater N in forested and low intensity (<10%) 

agriculture or urban land use areas is estimated to be 

fairly low, having 75th percentile concentrations of 0.5 

and 1.1 mg/L in two USGS studies (Nolan and Hitt, 

2002). They consequently concluded that a “reasonable” 

background concentration, as NO3-N, would be 1.1 mg/L, 

which would include effects in more sensitive aquifers 

with nominal loading for urban or agricultural sources. 

Nitrate can enter groundwater from a variety of 

sources, including all of those described in this report, 

but fertilizer and animal waste in rural, agricultural areas 

are especially prominent sources (Nolan and Ruddy, 

1996). Other sources include septic systems, more 

important in densely-developed and unsewered urban 

areas, and atmospheric deposition. Vulnerability to 

elevated NO3 levels is also variable, but an assessment 

and model by Nolan and Hitt (2006) predicted that “...

areas with high N application, high water input, welldrained soils, fractured rocks or those with high effective 

porosity, and lack of attenuation processes...” are 

especially vulnerable (r2 = 0.801).

As illustrated in Figure G-2, surveys confirm that 

NO3 in groundwater is elevated in many areas of the 

U.S., well above the 1.1 mg/L background upper bound 

described above (Figure G-2). The background colors 

on the map in Figure G-2 are indicative of different types 

of aquifers. In a 1992-1995 survey (Nolan and Stoner, 

2000), shallow groundwater underlying agricultural areas 

was found to be most severely impacted by elevated 

NO3-N levels (median concentration of 3.4 mg/L). 

Urban shallow aquifers were less impacted (median 

concentration of 1.6 mg NO3-N/L) and deeper, major 

Appendix G:

Impacts of Reactive Nitrogen on 

Aquatic Systems
112

aquifers, in general, had a median NO3-N concentration 

of 0.48 mg/L. However, NO3-N concentrations did 

exceed the 10 mg/L MCL threshold set by EPA for 

drinking water in more than 15% of the groundwater 

samples in the survey from drinking water aquifers. 

In the most recent survey of domestic well water 

quality (DeSimone et al., 2009), USGS found 

concentrations of NO3-N greater than 10 mg/L in 4.4% 

of the wells sampled. Concentrations exceeding the 

nitrate MCL were most frequently encountered in certain 

basins of the Southwest and California, west-central 

glacial aquifers in the Upper Midwest, and coastal plain 

aquifers and Piedmont crystalline rock aquifers in central 

Appalachia. Lowest concentrations were found in the 

coastal plain aquifers of the Southeast. In general, higher 

NO3-N concentrations were found near agricultural 

lands. In an additional analysis of shallow groundwater 

wells in agricultural areas, separate from the national 

survey, nearly 25% of the sampled wells exceeded the 10 

mg/L MCl for NO3-N. DeSimone et al. (2009) suggested 

that redox could be a defining factor in some cases, and 

perhaps was the reason for low NO3-N concentrations in 

the Southeast in soils that promote denitrification, as well 

as higher NO3-N levels in other areas where aquifers 

were better oxygenated.

Ammonia toxicity in freshwater systems

The EPA and states have long regulated ammonia 

(NH3) in the environment, not because of its nutrient 

contribution to cultural eutrophication, but because of 

its toxicity to freshwater aquatic life (U.S. EPA, 1986). 

The un-ionized ammonia molecule has been identified 

as the primary toxic form, rather than the ammonium 

ion (NH4

+), and research has further demonstrated 

the relationship between pH, temperature, and NH3

partitioning from the total ammonia pool in freshwaters 

(U.S. EPA, 1999). Because of the relationship to 

temperature in particular, and the variable sensitivity of 

species and life stages of aquatic organisms, state water 

quality standards in application generally consider cold 

and warm water conditions, as well as acute and chronic 

exposures to life stages of sensitive organisms, plus a 

margin of safety to derive criteria. 

While water quality criteria were initially set for 

concentrations of NH3, as criteria development skills 

and understanding improved, it made sense to develop 

criteria for total ammonia concentration for specific 

water quality conditions, e.g., cold or warm (salmonids 

present or absent) with consideration of ambient pH 

factors as appropriate, for protection of the most 

sensitive species likely to be present (early life stages 

present or absent), plus a margin of safety. This was 

because of evidence that the NH4

+ fraction may also 

be contributing to toxicity. Criteria could be presented 

as formulas in adopted state criteria to calculate total 

ammonia thresholds based on prevailing pH and 

temperature conditions and organisms present/absent, 

as appropriate.

Based on this research and analysis, EPA currently 

recommends adoption of ammonia criteria as criterion 

continuous concentration (CCC) and criterion maximum 

concentration (CMC) as described in Box G-1 (U.S. 

EPA, 1999).

Figure G-1: U.S. Population (1995) supplied by domestic drinking water wells 

Source: DeSimone et al., 2009.
113

Recent research on the sensitivity of freshwater 

unionid mussels (Family, Unionidae), and the rare and 

endangered status of several unionid species, has led 

the EPA to issue a draft update to the ammonia criteria 

guidance (U.S. EPA, 2009b) that would supersede 

the current 1999 guidance. Mussels have been found 

to be more sensitive to ammonia toxicity than the 

most sensitive species used to derive the 1999 criteria 

(Augspurger et al., 2003). Augspurger et al. (2003) found 

that the CMC with unionids considered would range 

from 1.75 to 2.50 mg total ammonia-N/L, 60% lower 

than the current method calculation of 5.62 mg total 

ammonia-N/L, for example. In the draft 2009 update, 

EPA proposed a two-tiered process for ammonia CMC 

and CCC development for waters with and without 

sensitive unionid species.

With this new sensitivity identified, ecosystem 

imbalances due to eutrophication and the presence and 

die off of invasive species may result in toxic levels of 

ammonia for mussels in freshwater systems, in addition 

to the conventional sewage and agricultural sources. For 

example, Cooper et al., (2005) suggested that die offs of 

invasive Asian clam (Corbicula fluminea), common in 

most southeastern U.S. waters, could produce enough 

sediment pore water ammonia during decay to be lethal 

to sensitive unionid mussels. Further, combination of 

ammonia with other toxic substances may compound 

toxic effects, as found by Wang et al. (2007a,b) for 

ammonia and copper. Their research suggests that the 

1999 criteria for total ammonia might not be protective 

of sensitive mussel species. 

Impacts of Nr on freshwater ecosystems

Reactive nitrogen (Nr), including reduced 

(ammonium, organic N compounds) and oxidized 

(nitrate, nitrite) forms, play central roles in modulating 

and controlling (limiting) primary and secondary 

production and species composition in freshwater 

ecosystems. These include lakes, reservoirs, streams, 

rivers, and wetlands (Goldman, 1981; Paerl, 1982; Elser 

et al., 1990, 2007; Wetzel, 2001). While phosphorus 

has been considered the primary limiting nutrient in 

freshwater ecosystems (c.f. Schindler, 1971; Schindler 

et al., 2008), there are numerous examples where Nr 

plays either a primary or secondary (i.e., co-limiting) 

role as a limiting nutrient (Paerl, 1982; North et al., 

2007; Wurtsbaugh et al., 1997; Lewis and Wurtsbaugh, 

2008). In particular, oligotrophic, alpine, tropical and 

subtropical, and other lakes having small watersheds 

Figure G-2: Nitrate concentrations in U.S. domestic drinking water wells 

(Background colors on the map are indicative of different aquifer types)

Source: DeSimone et al., 2009
114

relative to the lake surface/volume, and lakes 

experiencing incipient stages of eutrophication, tend to be 

N-limited (Wetzel, 2001; Lewis and Wurtsbaugh, 2008). 

N limitation was illustrated for Lake Tahoe (in California 

and Nevada) which was highly sensitive to N enrichment 

during its early stages of eutrophication (Goldman, 

1981; 1988). As the lake accumulated anthropogenic 

N inputs from both land-based runoff and atmospheric 

deposition within the Tahoe Basin, it began exhibiting 

symptoms of accelerating eutrophication, including 

noticeable “greening” of its formerly transparent nearshore waters and excessive epiphytic growth and fouling 

on its rocky bottom. Continued excessive N loading in 

the 1960s through 1980s has led to accelerating rates 

of algal primary production and a tendency to shift to 

more P limited conditions due to excessive N, relative 

to P, loading (Goldman,1988). This greater than 30-year 

progression to more eutrophic, and less desirable (from 

ecological, trophic and economic perspectives—i.e., 

tourism, water use) conditions has largely been spurred 

on by excessive N loading. Recent measures taken to 

reduce N inputs have been successful in reducing the 

lake’s rate of eutrophication (Goldman, 2002). Similarly, 

Lake Erie, which has experienced P-driven nuisance algal 

blooms starting in the 1950s, is now facing excessive N 

loading. This is largely a result of P input restrictions, 

which have been enacted since the 1970’s, accompanied 

by a lack of control on ever-increasing N loads. This 

shift in nutrient loading (increasing N:P) has led to a 

resurgence of toxic cyanobacterial blooms dominated 

by the non-N2 fixing genus Microcystis, an indicator of 

excessive N loading (North et al., 2007). 

Numerous lakes, reservoirs, rivers, estuaries (e.g., the 

Gulf of Mexico), and fjords worldwide exhibit N and 

Box G-1: The National Criterion for Ammonia in Fresh Water 

The available data for ammonia, evaluated using the procedures described in the “Guidelines for Deriving 

Numerical National Water Quality Criteria for the Protection of Aquatic Organisms and Their Uses,” indicate 

that, except possibly where an unusually sensitive species is important at a site, freshwater aquatic life should be 

protected if both of the following conditions are satisfied for the temperature (T) and pH of the waterbody:

1. The one-hour average concentration of total ammonia nitrogen (in mgN/L does not exceed, more than once every 

three years on the average, the CMC (acute criterion) calculated using the following equations. Where salmonid 

fish are present:

 Or where salmonid fish are not present:

2A. The 30-day average concentration of total ammonia nitrogen (in mg N/L) does not exceed, more than once 

every three years on the average, the CCC (chronic criterion) calculated using the following equations.

 When fish early life stages are present:

 When fish early life stages are absent:

2B. In addition, the highest 4-day average within the 30-day period should not exceed 2.5 times the CCC.

Source: U.S. EPA, 1999.
115

P co-limitation, either simultaneously or in seasonallyshifting patterns (Dodds et al., 1989; Elser et al., 1990, 

2007; Elmgren and Larsson, 2001; Forbes et al., 2008; 

Scott et al., 2008; Wetzel, 2001; North et al., 2007; Xu 

et al., 2010). For example, many reservoirs in Texas 

exhibit seasonal N limitation in the river-reservoir 

transition zone, regardless of their trophic status (Scott 

et al., 2009). Under these circumstances, N inputs tend 

to determine the spatial and temporal extent of summer 

nuisance algal blooms, a key symptom of degrading 

water quality (Dodds et al., 1989; Paerl, 2009; Xu et al., 

2010). N inputs, including those from increasing levels 

of atmospheric deposition, impact nutrient stoichiometry, 

with cascading effects on nutrient limitation, productivity, 

and lake nutrient cycling characteristics (Elser et al., 

2009). Therefore, the inputs of N play a critical role in the 

overall trophic response, trophic state, and water quality 

conditions of affected freshwater ecosystems.

In Florida lakes, algae are often limited by the 

availability of Nr (Kratzer and Brezonik, 1981). The 

most well-studied example is Lake Okeechobee, the 

largest lake in the Southeastern U.S., and a system that 

periodically displays large blooms of noxious blue-green 

algae. This lake has high availability of reactive P, and 

changes in the availability of Nr control the wax and 

wane of algae. In the 1980s and 1990s, blooms of algae 

were predominantly caused by cyanobacterial nitrogen 

(N2) fixer Anabaena. However, the most widespread 

recent bloom, which covered almost the entire lake 

surface in summer 2006, was caused by Microcystis, a 

non-N2- fixing cyanobacterium that depends on dissolved 

inorganic N (DIN: ammonium, nitrate, nitrite) and 

possibly organic N for its growth. This alga is the most 

common producer of toxins in Florida lakes, and it has 

the ability to “luxury consume” P from lake sediments 

and then rise through the water column, increasing 

its biomass to a level that largely is controlled by the 

amount of DIN. Because Lake Okeechobee’s sediments 

contain massive quantities of reactive P (Havens et al., 

2007), successful control of Microcystis blooms will 

require reduction in both P and N inputs to this lake.

In addition to the importance total N loads play 

in determining water quality status and trends, the 

supply rates and ratios of various Nr forms play an 

important role in structuring microalgal and macrophyte 

communities mediating freshwater primary production 

(Paerl, 1988; McCarthy et al., 2007, 2009; Lin et al., 

2008). For example, the ratio of ammonium to oxidized 

N was related to the proportion of cyanobacteria 

composing the total phytoplankton community of 

Lake Okeechobee (McCarthy et al., 2009). Non-N2-

fixing cyanobacteria, such as Microcystis, are superior 

competitors for reduced N (Blomqvist et al., 1994), 

but even N2-fixing cyanobacteria will preferentially 

assimilate ammonium if it is available (Ferber et al., 

2004). Ammonium is the initial N form produced by 

recycling processes (via invertebrate excretion and 

bacterial mineralization), but standing concentrations 

often remain very low because they are assimilated 

rapidly. Ammonium and other reduced N forms, such 

as dissolved free amino acids, are more available than 

oxidized N forms (nitrate and nitrite) to bacteria (Vallino 

et al., 1996) and cyanobacteria because less energy is 

required to incorporate reduced N into biomass than for 

oxidized forms (Syrett, 1981; Gardner et al., 2004; Flores 

and Herrero, 2005).

Lastly, it should be pointed out that both freshwater 

and marine systems do not respond to nutrient inputs 

in isolation. These systems are hydrologically and 

biogeochemically connected and coupled, functioning 

as a freshwater to marine continuum (Paerl, 2009). 

Nutrient limitation may shift along the continuum, and 

eutrophication and other symptoms of N and P overenrichment, including harmful algal blooms, hypoxia, 

loss of biodiversity, and food web alterations impact 

water quality, habitat condition, use, and sustainability 

of downstream waters. Therefore, excessive N loading 

in upstream freshwater ecosystems, ranging from the 

headwaters of pristine alpine streams to lowland lakes, 

reservoirs, and rivers can adversely affect downstream 

estuarine and coastal marine waters (Conley et al., 2009; 

Paerl, 2009). Examples of such continuum-scale impacts 

include such prominent systems as Chesapeake Bay, 

Albemarle-Pamlico Sound, Florida Bay, Mississippi 

River plume (Gulf of Mexico), Baltic Sea, and Coastal 

North Sea (Elmgren and Larsson, 2001; Boesch et al., 

2001; Paerl, 2009). 

Impact of nitrogen on wetlands

In this section, the possible impact of reactive nitrogen 

on wetlands is discussed. In Chapter 5 considerable 

attention is devoted to the subject of wetlands serving 

as effective nitrogen sinks. There are about 110 million 

ha of wetlands in the U.S., with more than half of 

those in the state of Alaska (Mitsch et al., 2009). Of 

those wetlands, roughly 97% are inland (and mostly 

freshwater) and 3% are estuarine (and mostly saline). Of 

the total wetlands, approximately half (166 million acres 

or 55 million ha) are peatlands, which, by their nature as 

low-nutrient systems, are most susceptible to nitrogen 

loadings, either from the atmosphere or from rivers and 

streams. More than any other ecosystem, wetlands are 

central to the cycling of nitrogen because they have both 

aerobic and anaerobic conditions that allow for a wide 

variety of important nitrogen processes, not the least of 

which is denitrification.

Wetlands are similar to lakes and streams and any 

other ecosystem in that their productivity is limited by 

nutrient availability. But with wetlands the hydrology 

limits or enhances productivity as well (Mitsch and 

Gosselink, 2007). The addition of excessive nutrients 

to wetlands, while often done purposefully when the 

wetlands are so-called treatment wetlands (Kadlec and 
116

Wallace, 2009), can cause vegetation shifts and decreases 

in plant diversity. Verhoeven et al. (2006) suggested 4.5 

g-N m-2 yr-1 as critical loading rate of nitrogen, generally 

from atmospheric sources, for peat-dominated wetlands. 

Morris (1991) had suggested that bogs and fens generally 

had loading rates of 1 to 6 g-N m-2 yr-1 respectively. 

Thus Verhoeven et al. (2006) were suggesting that 

wetlands should not be loaded beyond what is currently 

occurring in fen peatlands. These limitations do not 

apply for most mineral soil wetlands, particularly those 

connected to streams and rivers. Most freshwater and 

tidal marshes have nitrogen loading rates closer to 60 

g-N m-2 yr-1 and they maintain a reasonably high and 

sustainable productivity. Using a sustainable rate of 

nitrogen retention as a measure, Mitsch and Jørgensen 

(2004) suggest a range of nitrogen retention rate of 10-20 

g-N m-2 yr-1 for wetlands to maintain their biodiversity 

while being nitrogen sinks at the same time. Overall, this 

is a fruitful direction for wetland research to determine 

the assimilative capacity of wetlands for nutrients, 

including nitrogen, while not surpassing a limit that will 

dramatically change the wetland’s structure and function.

Impacts of Nr on coastal systems

Mitsch et al. (2001) suggest that streams and rivers 

themselves are not always as much affected by nutrient 

loading as are lakes, wetlands, coastal areas, and other 

lentic bodies of water. However, in most cases, these 

nutrient-enriched waterways flow to the sea, with 

eutrophication of coastal waters the unfortunate result. 

This problem now occurs regularly throughout the world 

(World Resources Institute, 2008), in locations such as 

the Gulf of Mexico (Rabalais et al., 1996), the Baltic Sea 

(Larson et al., 1985), and the Black Sea (Tolmazin, 1985).

During the past century, following large-scale use of 

synthetic N fertilizers in agriculture, rapid expansion 

of industrial and transportation-related fossil fuel 

combustion and coastal urbanization, humans have 

significantly altered the balance between “new” N inputs 

and N losses in the marine environment (Codispoti et al., 

2001; Galloway and Cowling, 2002). During this time 

frame, terrestrial discharge and atmospheric N emissions 

have increased 10-fold (Howarth et al., 1996; Holland 

et al., 1999). This number keeps growing as human 

development continues to expand in coastal watersheds 

(Vitousek et al., 1997a,b).

Researchers have long recognized this growing 

imbalance, especially in estuarine and coastal waters 

where anthropogenically-derived N over-enrichment 

has fueled accelerated primary production, or “cultural” 

eutrophication (Vollenweider et al., 1992; Nixon, 1995). 

Eutrophication is a condition where nutrient-enhanced 

primary production exceeds the ability of higher ranked 

consumers and organic matter-degrading microbes to 

consume and process it. D’Elia (1987) characterized 

this condition as “too much of good thing” or overfertilization of N-limited marine ecosystems with “new” 

N, the bulk of it being anthropogenic (Howarth et al., 

1996; Vitousek et al., 1997a,b; Galloway and Cowling, 

2002). Symptoms of N-driven eutrophication vary: from 

subtle increases in plant production to changes in primary 

producer community composition; to rapidly accelerating 

algal growth, visible discoloration or blooms, losses in 

water clarity, increased consumption of oxygen, dissolved 

oxygen depletion (hypoxia), which is stressful to resident 

fauna and flora; to, in the case of total dissolved oxygen 

depletion (anoxia), elimination of habitats (Paerl, 

1988, 1997; Diaz and Rosenberg, 1995; Rabalais and 

Turner, 2001). Other effects include submerged aquatic 

vegetation (SAV) losses, possible impacts on tidal 

wetland health, and disruption of estuarine food chain 

dynamics that may favor an imbalance towards lower 

trophic levels (e.g., jellyfish).

Anthropogenic or cultural eutrophication has 

been closely linked to population densities in coastal 

watersheds (Peierls et al., 1991; Nixon, 1995; Vitousek 

et al., 1997a,b). Primary sources of N enrichment 

include urban and agricultural land uses as well as 

wastewater treatment plants, many of which have not 

been designed to remove N. A significant, and in many 

instances increasing, proportion of “new” N input can 

also be attributed to remote sources residing in airsheds. 

Delivery routes can also be complex, especially when 

via subsurface aquifers outside the immediate watershed, 

which can confound source definition and create long 

delays in delivery and management response (Paerl, 1997; 

Jaworski et al., 1997; Galloway and Cowling, 2002; Paerl 

et al., 2002). 

The availability of N controls primary production in 

much of the world’s estuarine, near-shore coastal, and 

open-ocean waters (Dugdale, 1967; Ryther and Dunstan, 

1971; Nixon, 1995; Paerl, 1997; Boesch et al., 2001). 

As previously discussed, nitrogen can also play a role 

as either a primary or secondary limiting nutrient in 

freshwater environments, especially large lakes (e.g., 

Lake Tahoe, Lake Superior). As such, the fertility of 

these waters is often closely controlled by N inputs, 

which are provided either internally by regeneration of 

pre-existing N and biologically-fixed atmospheric N2, or 

supplied externally (i.e. “new” N) as combined N sources 

delivered via surface runoff, sub-surface groundwater, or 

atmospheric deposition. 

The extent to which accelerated N loading promotes 

eutrophication and its symptoms varies greatly among 

marine ecosystems. Receiving waters exhibit variable 

sensitivities to N and other nutrient [phosphorus (P), iron 

(Fe), and silica (Si)] loads that are controlled by their size, 

hydrologic properties (e.g., flushing rates and residence 

times), morphologies (depth, volume), vertical mixing 

characteristics, geographic and climatic regimes and 

conditions. In addition, the magnitude and distribution of 

N in relation to other nutrient loads can vary substantially. 

In waters receiving very high N loads relative to 
117

requirements for sustaining primary and secondary 

production, other nutrient limitations may develop. This 

appears to be the case in coastal waters downstream of 

rivers draining agricultural regions that are enriched in N. 

On the ecosystem level, estuarine and coastal waters 

exhibit individualistic responses to N loads over seasonal 

and longer (multi-annual, decadal) time scales. The 

degree to which these systems are exposed to freshwater 

discharge, tidal exchange, and vertical mixing is critical 

for determining how they respond to specific N loads 

(Vollenweider et al., 1992; Nixon, 1995; Cloern, 1999, 

2001; Valdes-Weaver et al., 2006; Paerl et al., 2007). 

Another variable is the manner in which N loading takes 

place, which may range from acute pulsed events such 

as storms and associated flooding, to longer-term gradual 

(chronic) increases in N loading associated with more 

predictive seasonal, annual and inter-annual hydrologic 

cycles. There are striking contrasts in ecosystem response 

to N inputs that reflect a range in physical (hydrodynamic, 

optical) and climatic conditions (Cloern, 1999, 2001). 

Examples include contrasts between strong tidallydriven estuarine systems, such as Delaware Bay and San 

Francisco Bay, and non-tidal, lagoonal systems, such as 

North Carolina’s Pamlico Sound and Texas’s Luguna 

Madre, or semi enclosed coastal systems, such as Florida 

Bay and the Long Island Sound (Bricker et al., 1999; 

Valdes-Weaver et al., 2006; Paerl et al., 2007). 

Externally-supplied N comes in various forms, 

including organic N and inorganic reduced (NH3 and 

NH4

+ ion) and oxidized (NO3–) N, all of which are 

potentially available to support new production and 

eutrophication. Laboratory experiments on phytoplankton 

isolates and bioassays with natural phytoplankton 

communities have indicated that these contrasting 

forms may be differentially and preferentially utilized, 

indicating that, depending on composition of the affected 

phytoplankton community, some forms are more reactive 

than others (Collos, 1989; Stolte et al., 1994; Riegman, 

1998). Phytoplankton community composition can also 

be altered by varying proportions and supply rates of 

different forms of N (Dortch, 1990; Stolte et al., 1994; 

Harrington, 1999; Pinckney et al., 1999; Piehler et al., 

2002). Monitoring and research on dissolved organic N 

inputs and their effects should be conducted in receiving 

streams, rivers, lakes, estuarine, and coastal waters, since 

there is evidence that these compounds can be utilized by 

phytoplankton, including harmful bloom species (Paerl, 

1988; Antia et al., 1991; Carlsson and Granéli, 1998; 

Gilbert et al., 2006). In addition, specific N compounds 

may interact with light availability, hydrodynamics and 

other nutrients, most notably P, Si, Fe, and trace metals, 

to influence phytoplankton community growth rates and 

composition (Harrison and Turpin, 1982; Smith, 1990; 

Dortch and Whitledge, 1992). 

Over the past 25 years, there has been a growing 

recognition of cultural eutrophication as a serious 

problem in coastal estuaries (NRC, 2000). Globally, 

Selman et al. (2008) have reported “Of the 415 areas 

around the world identified as experiencing some form 

of eutrophication, 169 are hypoxic and only 13 systems 

are classified as ‘systems in recovery.”Comprehensive 

surveys of U.S. estuaries have been conducted by the 

National Oceanic and Atmospheric Administration 

(NOAA) as part of the National Estuarine Eutrophication 

Assessments (NEEA) in 1999 and 2004 (Bricker et al., 

1999, 2007). The most recent report, released in 2007 

(Bricker et al., 2007) focused on nutrient enrichment 

and its manifestations in the estuarine environment and 

relies on participation and interviews of local experts 

to provide data for the assessment. Among the key 

findings for nearly 100 assessed U.S. estuaries were that 

eutrophication is a widespread problem, with the majority 

of assessed estuaries showing signs of eutrophication— 

65% of the assessed systems, representing 78% of 

assessed estuarine area, had moderate to high overall 

eutrophic conditions. The most common symptoms of 

eutrophication were high spatial coverage and frequency 

of elevated chlorophyll a (phytoplankton)—50% of the 

assessed estuaries, representing 72% of assessed area, had 

a high chlorophyll a rating. 

Further field evaluations by EPA and state and 

university collaborators under the National Coastal 

Assessment (NCA) used probabilistic monitoring 

techniques. The NCA National Coastal Condition Reports 

(NCCR) (U.S. EPA, 2001a, 2004, 2006b) are more closely 

related to nutrient enrichment assessments, especially for 

manifestations of nutrient enrichment such as hypoxia, 

nuisance algal blooms, and general habitat degradation. 

The last comprehensive national NCCR was published 

in 2004 (U.S. EPA, 2004) with a more recent assessment 

focused on 28 National Estuary Program estuaries 

published in 2007 (U.S. EPA, 2006). The 2004 NCCR 

included an overall rating of “fair” for estuaries, including 

the Great Lakes, based on evaluation of more than 2,000 

sites. The water quality index, which incorporates nutrient 

effects primarily as chlorophyll-a and dissolved oxygen 

impacts, was also rated “fair” nationally. Forty percent 

of the sites were rated “good” for overall water quality, 

while 11% were “poor” and 49% “fair.” 

Attainment of water quality management goals 

and standards for coastal systems

Estuarine systems, where bio-available Nr is 

more likely to be the limiting nutrient, are most often 

susceptible to Nr enrichment (Paerl, 1997; Boesch et al., 

2001). Defining single-number criteria for nutrients or 

related indicators representative of undesirable levels of 

productivity (e.g., chlorophyll-a) is difficult, even using 

the ecoregional approach recommended by EPA. State 

managers more often use the formal TMDL process or 

collaborative estuarine management plans to set site- or 

estuary-specific N management targets to meet existing, 

related water quality criteria (e.g., dissolved O2 or 
118

chlorophyll a). Some of the more prominent efforts and 

targets for nitrogen control are summarized in Table G-1.

These targets all exceed nitrogen load reduction goals 

that the Committee has found to be readily achievable 

using existing technology and management authority 

(i.e., less than 25% from specific source categories). 

Some sources of nitrogen loading to the estuaries in 

Table G-1 pose greater management challenges and 

the expectation is that reductions in Nr loadings to 

estuaries would cumulatively be less than 25%. This 

suggests that efforts will be needed to enable even 

greater nitrogen load reductions in the future. Many of 

the management actions the Committee has proposed in 

this report would require substantive changes in national 

programs, regulatory authority, management technologies 

and societal demands to be accomplished. This is a 

nutrient management concern that state managers are 

well aware of as they develop TMDLs and management 

plans that range above attainment potential, not only 

for Nr but more frequently for other pollutants that are 

predominately nonpoint source and stormwater loaded 

(including atmospheric source contributions).

The Chesapeake Bay Program, for example, is 

a model for Nr and P management in many ways. 

Considerable resources were committed, and many 

BMPs were implemented. Yet despite regional efforts 

and commitments from all watershed states, and more 

funding than any other estuary program is likely 

to see, management targets have not been met, and 

recent data (2007) reveal the occurrence of a severe 

hypoxic episode. Concerns over the slow progress in 

restoring the Chesapeake Bay led to the issuance of an 

Executive Order on May 15, 2009, establishing a Federal 

Leadership Committee led by the EPA to develop and 

implement a plan to restore the Bay in collaboration with 

state agencies (Federal Register 74(93): 23097-23104). 

Similarly, the adoption of the Long Island Sound TMDL, 

which was driven by the presence of reactive nitrogen 

(see Box G-2), sets an implementation plan that could 

attain Connecticut and New York dissolved oxygen 

criteria, but only if “alternative technologies” such as 

mechanical aeration of the Sound or biological harvesting 

of nutrients, are used.

Table G-1: Estuaries with nitrogen management plans or 

TMDLs and percent nitrogen load reduction targets

Estuary Nitrogen Load 

Reduction Target TMDL or Plan

Casco Bay, Maine 45% Plan

Chesapeake Bay >40% Plan

Northern Gulf of Mexico

Mississippi Plume Region 45% Plan

Long Island Sound 60% for CT & NY sources TMDL

Neuse River Estuary, NC 30% TMDL

Tampa Bay, FL Maintain TN (total nitrogen) 

load at 1992-1994 levels TMDL & Plan
119

Box G-2: Long Island Sound Total Maximum Daily Load: Focus on Reactive Nitrogen

A TMDL sets a goal for reducing the load of a specific pollutant that is causing impairment to a waterbody. In 

the case of Long Island Sound, the impairment constitutes low concentrations of dissolved O2 that violate both 

Connecticut’s and New York’s water quality standards. Nitrogen has been identified as the pollutant that causes 

substandard levels of dissolved oxygen in Long Island Sound and, accordingly, Connecticut’s and New York’s 

environmental agencies have developed a TMDL that assigns nitrogen reductions from both point sources (the 

wasteload allocation or WLA) and nonpoint sources (the load allocation or LA) in their respective states to meet the 

established 58.5% reduction of anthropogenic sources.

The Long Island Sound TMDL is set at 23,966 tons of N/year, which represents a 23,834 ton/year reduction from 

the total baseline (anthropogenic + natural sources considered) of 47,788 tons/year from Connecticut and New 

York only. Most of that N load comes from point sources – POTWs (publicly owned treatment works) and CSOs 

(combined sewer overflows) – accounting for 38,899 tons/yr of the total N load from the two states, or 81% of the 

load. For that reason, the focus has been on managing point sources, although attainment of water quality standards 

will require more widespread reductions from atmospheric deposition, stormwater, and nonpoint sources, and from 

other watershed states north of Connecticut.

Connecticut and New York have some flexibility in the apportionment of those reductions between the WLA and the 

LA, but must have completed 40% of the required reductions by 2004, 75% by 2009, and 100% by 2014 when the 

final TMDL will be met. However, the TMDL is presently undergoing revision to incorporate findings from a new 

model of Long Island Sound, and to reflect changes in dissolved O2 criteria in both states. The revised TMDL will 

likely require more aggressive reductions of nitrogen to meet dissolved O2 criteria and may formalize targets for 

upstream state contributions and atmospheric deposition.
120
121

There are limits to how much plant growth can be 

increased by N fertilization. At some point, when the 

natural N deficiencies in an ecosystem are fully relieved, 

plant growth becomes limited by availability of other 

resources such as phosphorus, calcium, or water and the 

vegetation can no longer respond to further additions of 

Nr. In theory, when an ecosystem is fully Nr-saturated 

and its soils, plants, and microbes cannot use or retain 

any more, all new Nr deposits will be dispersed to 

streams, groundwater, and the atmosphere. Nr saturation 

has a number of damaging consequences for the health 

and functioning of ecosystems. These impacts first 

became apparent in Europe almost three decades ago 

when scientists observed significant increases in nitrate 

concentrations in some lakes and streams and also 

extensive yellowing and loss of needles in spruce and 

other conifer forests subjected to heavy Nr deposition. In 

soils, most notably forest soils because of their natural 

low pH, as NH4

+ builds up it is converted to nitrate by 

bacterial action, a process that releases hydrogen ions 

and contributes to soil acidification. The buildup of NO3ˉ 

enhances emissions of nitrous oxides from the soil and 

also encourages leaching of highly water-soluble NO3ˉ 

into streams or groundwater. As negatively charged NO3ˉ 

seeps away, positively charged alkaline minerals such as 

calcium, magnesium, and potassium are carried along. 

Thus, soil fertility is decreased by greatly accelerating 

the loss of calcium and other nutrients that are vital 

for plant growth. As calcium is depleted and the soil 

acidified, aluminum ions are mobilized, eventually 

reaching toxic concentrations that can damage tree roots 

or kill fish if the aluminum washes into streams (Vitousek 

et al., 1997a,b).

Forests, grasslands, and wetlands vary substantially 

in their capacity to retain added nitrogen. Interacting 

factors that are known to affect this capacity include 

soil texture, degree of chemical weathering of soil, 

fire history, rate at which plant material accumulates, 

and past human land use. However, we still lack a 

fundamental understanding of how and why N-retention 

processes vary among ecosystems, much less how they 

have changed and will change with time and climate 

change (Clark and Tilman, 2008).

An overarching impact of excess Nr on unmanaged 

terrestrial ecosystems is biodiversity loss. In North 

America, dramatic reductions in biodiversity have been 

created by fertilization of grasslands in Minnesota 

and California. In England, N fertilizers applied to 

experimental grasslands have led to similarly increased 

dominance by a few N-responsive grasses and loss 

of many other plant species. In formerly species-rich 

heathlands across Western Europe, Nr deposition has 

been blamed for great losses of biodiversity in recent 

decades, with shallow soils containing few alkaline 

minerals to buffer acidification (Vitousek et al., 1997a,b; 

Bobbink et al., 2010).

Losses of biodiversity driven by Nr deposition can in 

turn affect other ecological processes. Experiments in 

Minnesota grasslands showed that in ecosystems made 

species-poor by fertilization, plant productivity was much 

less stable in the face of a major drought. Even in nondrought years, the normal vagaries of climate produced 

much more year-to-year variation in the productivity of 

species-poor grassland plots than in more diverse plots 

(Vitousek et al., 1997a,b).

Appendix H:

Nr Saturation and Ecosystem Function
122
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