Science Advisory Board Office of the Administrator 

# Reactive 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 Repo r t of the EP A Science Advi s o r y Bo ar dC over photos by E ric V ance and other photographers, U.S. EP A, courtesy of U.S. EP A.i T he H onorable Lisa P . Jackson Administrator U.S. E nvironmental P rotection Agency 1200 P ennsylvania Avenue, N .W

. Washington, D

. C . 20460

Subject: Reactive N itrogen in the United States: An Analysis of I nputs, F lows, C onsequences, and Management O ptions – A Report of the EP A Science Advisory Board

D ear Administrator Jackson:

E xcess 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. I n addition, reactive nitrogen is associated with harmful human health ef fects caused by air pollution and drinking water contamination. Reactive nitrogen (hereafter referred to as N r) 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 N 2
. EP A and other federal and state agencies have implemented programs to reduce the risks posed by excessive N r , but a more comprehensive and integrated approach is needed to manage the use of N r 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 dif ferent forms and media.

T he Science Advisory Board (SAB) I ntegrated N itrogen C ommittee has conducted a study to analyze sources and fate of N r in the United States and provide advice to EP A 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. O ur objectives for this study were to:

• I dentify and analyze from a scientific perspective the problems N r presents in the environment and the links among

them;

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

• I dentify additional risk management options for EP A’ s consideration; and

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

T he 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.

I n general, the SAB finds that:

• I n the United States, human activities across multiple sources currently introduce more than five times the N r into

the environment than natural processes. T he lar gest U.S. sources of new N r entering the U.S. environment include: the creation and use of synthetic fertilizers, N r created by legumes, and the combustion of fossil fuels.

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

N r formed during fossil fuel combustion.

• T he introduction of human created N r 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 A D M INI S T RA TO R
S CIENCE A DVI S O R Y

B O AR D EP A-SAB-11-013ii • Multiple strategies and actions exist to more effectively minimize the inputs of N r to the environment and maximize nitrogen use efficiency.

T he SAB provides the following overarching recommendations to improve the management of N r.

• T he 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 N r are required to address the significant environmental and human health damage caused by N r. N ew institutional structures and relationships that also reflect the multi-media and multi-form character of N r and its flows and transformations through the environment will have to be created for ef fective control and management.

• T he SAB recommends an integrated approach to the management of N r. T his 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 N r and on decreasing the risks of N r. T he 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 ef fectiveness.

• EP A should form an intra-Agency N r management task force that will build on the existing breadth of N r research and management capabilities within the Agency

. I ts objective should be to increase scientific understanding of: (1) N r impacts on terrestrial and aquatic ecosystems, human health, and climate; (2) N r-relevant monitoring requirements; and (3) the most efficient and cost effective means by which to decrease various adverse impacts of N r loads as they cascade through the environment.

• Successful N r management will require changes in the way EP A interacts with other agencies. T he SAB recommends that EP A convene a reactive nitrogen inter-agency management task force with broad representation from other agencies and departments involved with N r control or utilization. T his is essential to coordinate federal programs that address N r concerns and would help ensure clear responsibilities for monitoring, modeling, researching, and managing N r 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.

I n the context of addressing the specific study objectives, the SAB explored how an estimated 25 percent reduction in N r introduced into the environment might be achieved with existing technology in the coming 10 to 20 years through actions that could be taken by EP A, other management authorities, and other public and private or ganizations. 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 N r in surface waters, decreased loss of reactive nitrogen from agricultural lands and animal feeding operations, and decreased dischar ge of reactive nitrogen from point sources and developed (urban) lands. H owever, dealing ef fectively with reactive nitrogen’scascade through air, water and land will require an integrated management approach that is multi-media and multistressor as suggested by recent initiatives by EP A’s O ffice of Research and D evelopment.

T he most important task for EP A and allied agencies and departments will be to effectively inform the public of the costs and dangers of excess N r. Without strong public support, the widespread efforts necessary to control N r will not be possible.

I n closing, we appreciate the opportunity to provide advice on this very important topic, and we look forward to receiving your response. T he SAB stands ready to provide more information as it may be useful and would be pleased to assist EP A in the implementation of the report’s recommendations, if the EP A would find such support valuable. Sincerely,

D r. O tto C

. D oering III C hair I ntegrated N itrogen C ommittee Science Advisory Board D r. James N . Galloway C hair (2007-2009) I ntegrated N itrogen C ommittee Science Advisory Board D r. T homas L. T heis V

ice- C hair (2008-2009) I ntegrated N itrogen C ommittee Science Advisory Board

D r. D ebo rah L. Swackhamer C hair Science Advisory Boardiii CHAIR Dr. Otto C. Doering III
, P rofessor
, D epartment of Agricultural E conomics, P urdue University , W. Lafayette, IN CHAIR (2007-2009) Dr. James N. Galloway , Sidman P

. P oole P rofessor of E nvironmental Sciences, Associate D ean for the Sciences, C ollege and Graduate School of Arts and Sciences, University of V irginia, C harlottesville, V A VICE-CHAIR (2008-2009) Dr. Thomas L. Theis
, D irector
, I nstitute for E nvironmental Science and P olicy , University of I llinois at C hicago, C hicago, I L

MEMBERS Dr . Viney Aneja
, P rofessor
, D epartment of Marine, E arth, and Atmospheric Sciences, School of P hysical and Mathematical Sciences, N orth C arolina State University , Raleigh, NC
Dr. Elizabeth Boyer , Associate P rofessor , School of F orest Resources and Assistant D irector
, P ennsylvania State I nstitutes of E ner gy & the E nvironment, and D irector
, P ennsylvania Water Resources Research C enter , P ennsylvania State University , University P ark, P A

Dr. Kenneth G. Cassman
, P rofessor
, D epartment of Agronomy and H orticulture, I nstitute of Agriculture and N atural Resources, University of N ebraska, Lincoln, NE
Dr. Ellis B. Cowling , University D istinguished P rofessor At-Large E meritus, C olleges of N atural Resources and Agriculture and Life Sciences, N orth C arolina State University , Raleigh, NC

Dr. Russell R. Dickerson
, P rofessor
, D epartment of Atmospheric and O ceanic Science, T he University of Maryland, C ollege P ark, M D

Mr. William Herz
, V ice P resident for Scientific P rograms, T he F ertilizer I nstitute, Washington, DC
Dr. Donald L. Hey
, E xecutive D irector , Wetlands Research, I nc, C hicago, I L

Dr . Richard Kohn
, P rofessor , Animal Sciences D epartment, University of Maryland, C ollege P ark, M D
Dr. JoAnn S. Lighty
, C hair and P rofessor
, C hemical E ngineering, University of Utah, Salt Lake C ity , U T

Dr. William Mitsch
, P rofessor
, O lentangy River Wetland Research P ark, T he O hio State University
, C olumbus, OH

Dr. William Moomaw
, P rofessor of I nternational E nvironmental P olicy and D irector of the C enter for I nternational E nvironment and Resource P olicy
, T he F letcher School of Law and D iplomacy
, T ufts University, Medford, MA

Dr . Arvin Mosier
, V isiting P rofessor , Agricultural and Biological E ngineering D epartment, University of F lorida, Mount P leasant, S C

Dr. Hans Paerl
, P rofessor of Marine and E nvironmental Sciences, I nstitute of Marine Sciences, University of N orth C arolina – C hapel H ill, Morehead C ity
, NC
Dr. Bryan Shaw
, C ommissioner
, T exas C ommission on E nvironmental Quality , Austin, T X

Mr . Paul Stacey
, D irector , Bureau of Water Management and Land Reuse, P lanning and Standards D ivision, C onnecticut D epartment of E nvironmental P rotection, H artford, CT 

# U.S. Environmental Protection Agency Science Advisory Board

Integrated Nitrogen Committeeiv SCIENCE ADVISORY BOARD STAFF Dr. Thomas Armitage
, D esignated F ederal O fficer (2009-present), U.S. E nvironmental P rotection Agency, Washington, DC

Dr. Angela Nugent
, D esignated F ederal O fficer (2009), U.S. E nvironmental P rotection Agency, Washington, DC

Ms. Kathleen White
, D esignated F ederal O fficer (2007-2008), U.S. E nvironmental P rotection Agency, Washington, DCv CHAIR Dr. Deborah L. Swackhamer
, P rofessor and C harles M. D enny , Jr., C hair in Science, T echnology and P ublic P olicy and C o- D irector of the Water Resources C enter , H ubert H

. H umphrey I nstitute of P ublic Affairs, University of Minnesota, St. P aul, M N

SAB MEMBERS Dr. David T. Allen
, P rofessor
, D epartment of C hemical E ngineering, University of T exas, Austin, T X

Dr . Claudia Benitez-Nelson
, F ull P rofessor and D irector of the Marine Science P rogram, D epartment of E arth and O cean Sciences, University of South C arolina, C olumbia, S C

Dr. Timothy Buckley , Associate P rofessor and C hair
, D ivision of E nvironmental H ealth Sciences, C ollege of P ublic H ealth, T he O hio State University
, C olumbus, OH

Dr. Patricia Buffler
, P rofessor of E pidemiology and D ean E merita, D epartment of E pidemiology , School of P ublic H ealth, University of C alifornia, Berkeley
, C A

Dr . Ingrid Burke
, D irector
, H aub School and Ruckelshaus I nstitute of E nvironment and N atural Resources, University of Wyoming, Laramie, W Y

Dr. Thomas Burke
, P rofessor
, D epartment of H ealth P olicy and Management, Johns H opkins Bloomber g School of P ublic H ealth, Johns H opkins University , Baltimore, M D

Dr. Terry Daniel
, P rofessor of P sychology and N atural Resources, D epartment of P sychology , School of N atural Resources, University of Arizona, T ucson, AZ

Dr . George Daston
, V ictor Mills Society Research F ellow
, P roduct Safety and Regulatory Affairs, P rocter & Gamble, C incinnati, OH

Dr. Costel Denson , Managing Member, C ostech T echnologies, LL C
, N ewark, DE
Dr. Otto C. Doering III
, P rofessor
, D epartment of Agricultural E conomics, P urdue University , W. Lafayette, IN

Dr. David A. Dzombak , Walter J. Blenko Sr. P rofessor of E nvironmental E ngineering, D epartment of C ivil and E nvironmental E ngineering, C ollege of E ngineering, C arnegie Mellon University
, P ittsbur gh, P A

Dr. T. Taylor Eighmy
, V ice P resident for Research, O ffice of the V ice P resident for Research, T exas T ech University, Lubbock, T X

Dr . Elaine Faustman
, P rofessor and D irector
, I nstitute for Risk Analysis and Risk C ommunication, School of P ublic H ealth, University of Washington, Seattle, WA
Dr. John P. Giesy
, P rofessor and C anada Research C hair
, V eterinary Biomedical Sciences and T oxicology C entre, University of Saskatchewan, Saskatoon, Saskatchewan, C anada

Dr . Jeffrey K. Griffiths
, P rofessor
, D epartment of P ublic H ealth and C ommunity Medicine, School of Medicine, T ufts University, Boston, MA

Dr . James K. Hammitt
, P rofessor
, C enter for Risk Analysis, H arvard University , Boston, MA

Dr. Bernd Kahn
, P rofessor E meritus and Associate D irector
, E nvironmental Radiation C enter , Georgia I nstitute of T

echnology, Atlanta, GA
Dr. Agnes Kane
, P rofessor and C hair
, D epartment of P athology and Laboratory Medicine, Brown University
, P rovidence, R I

Dr. Madhu Khanna
, P rofessor
, D epartment of Agricultural and C onsumer E conomics, University of I llinois at Urbana- C hampaign, Urbana, I L

Dr . Nancy K. Kim , Senior E xecutive, H ealth Research, I nc., T roy, NY 

# U.S. Environmental Protection Agency Science Advisory Boardvi Dr. Catherine Kling
, P rofessor
, D epartment of E conomics, I owa State University , Ames, I A

Dr . Kai Lee
, P rogram O fficer
, C onservation and Science P rogram, D avid & Lucile P ackard F oundation, Los Altos, C A

Dr . Cecil Lue-Hing, President
, C ecil Lue- H ing & Associates I nc., Burr Ridge, I L

Dr . Floyd Malveaux
, E xecutive D irector , Merck C hildhood Asthma N etwork, I nc., Washington, DC

Dr. Lee D. McMullen
, Water Resources P ractice Leader , Snyder & Associates, I nc., Ankeny, I A

Dr . Judith L. Meyer
, P rofessor E meritus, O dum School of E cology , University of Georgia, Lopez I sland, WA

Dr. James R. Mihelcic
, P rofessor
, C ivil and E nvironmental E ngineering, State of F lorida 21st C entury World C lass Scholar , University of South F lorida, T ampa, F L

Dr . Jana Milford
, P rofessor
, D epartment of Mechanical E ngineering, University of C olorado, Boulder
, CO

Dr. Christine Moe
, E ugene J. Gangarosa P rofessor
, H ubert D epartment of Global H ealth, Rollins School of P ublic H ealth, E mory University , Atlanta, GA
Dr. Horace Moo-Young
, D ean and P rofessor
, C ollege of E ngineering, C omputer Science, and T echnology, C alifornia State University , Los Angeles, C A

Dr . Eileen Murphy , Grants F acilitator
, E rnest Mario School of P harmacy , Rutgers University, P iscataway
, N J

Dr . Duncan Patten , Research P rofessor
, H ydroecology Research P rogram, D epartment of Land Resources and E nvironmental Sciences, Montana State University , Bozeman, M T
Dr. Stephen Polasky
, F esler -Lampert P rofessor of E cological/ E nvironmental E conomics, D epartment of Applied E conomics, University of Minnesota, St. P aul, M N

Dr. C. Arden Pope III
, P rofessor
, D epartment of E conomics, Brigham Y oung University, P rovo, U T
Dr. Stephen M. Roberts
, P rofessor
, D epartment of P hysiological Sciences, D irector
, C enter for E nvironmental and H uman T oxicology, University of F lorida, Gainesville, F L

Dr . Amanda Rodewald
, P rofessor of Wildlife E cology , School of E nvironment and N atural Resources, T he O hio State University
, C olumbus, OH

Dr. Jonathan M. Samet
, P rofessor and F lora L. T hornton C hair
, D epartment of P reventive Medicine, University of Southern C alifornia, Los Angeles, C A

Dr . James Sanders
, D irector and P rofessor , Skidaway I nstitute of O ceanography , Savannah, GA

Dr. Jerald Schnoor , Allen S. H enry C hair P rofessor
, D epartment of C ivil and E nvironmental E ngineering, C o- D irector
, C enter for Global and Regional E nvironmental Research, University of I owa, I owa C ity
, I A

Dr . Kathleen Segerson
, P hilip E

. Austin P rofessor of E conomics, D epartment of E conomics, University of C onnecticut, Storrs, CT

Dr. Herman Taylor
, D irector
, P rincipal I nvestigator , Jackson H eart Study , University of Mississippi Medical C enter , Jackson, MS
Dr. Barton H. (Buzz) Thompson, Jr. , Robert E

. P aradise P rofessor of N atural Resources Law at the Stanford Law School and P erry L. Mc C arty D irector , Woods I nstitute for the E nvironment, Stanford University , Stanford, C A

Dr . Paige Tolbert
, P rofessor and C hair
, D epartment of E nvironmental H ealth, Rollins School of P ublic H ealth, E mory University , Atlanta, GA

Dr. John Vena
, P rofessor and D epartment H ead, D epartment of E pidemiology and Biostatistics, C ollege of P ublic H ealth, University of Geor gia, Athens, GA

Dr. Thomas S. Wallsten
, P rofessor and C hair
, D epartment of P sychology , University of Maryland, C ollege P ark, M D

Dr. Robert Watts
, P rofessor of Mechanical E ngineering E meritus, T ulane University, Annapolis, M D

Dr. R. Thomas Zoeller
, P rofessor
, D epartment of Biology , University of Massachusetts, Amherst, MA

SCIENCE ADVISORY BOARD STAFF Dr. Angela Nugent
, D esignated F ederal O fficer , U.S. E nvironmental P rotection Agency, Washington, DCvii 

# Acknowledgements

T he EP A Science Advisory Board I ntegrated N itrogen C ommittee would like to acknowledge many individuals who provided their perspectives and insights for the C ommittee’ s consideration in the development of this report.

I nvited speakers provided perspectives at public meetings of the C ommittee. T hese speakers included: the following individuals from the U.S. EP A - Mr . Robert Bastian, Mr. Gilbert C astellanos, Mr . John D avies, D r . Robin D ennis, D r

. Jonathan Garber, D r . Richard H aeuber
, D r . Alan H echt, Mr . Rudolph Kapichak, Mr. Gary Lear, D r . Richard Linthurst, D r . Rohit Mathur, Ms. Roberta P arry , and Mr. James P ender gast; the following individuals from the US D A - D r . Roger C laassen, D r . Stan D aberkow
, D r . Raymond Knighton, and D r . Mark Walbridge; and the following individuals from other organizations - Mr. C raig C ox, Soil and Water C onservation Society; D r . Jan Willem E risman, E ner gy Research C entre of the N etherlands; D r

. P aul F ixen, I nternational P lant N utrition I nstitute; D r

. D avid Mc N aught, E nvironmental D efense F und; D r . John Miranowski, I owa State University; D r . Martin P etrovic, C ornell University; and Mr . John Sheehan, Live F uels.

T he C ommittee would also like to thank the speakers and participants who attended the SAB I ntegrated N itrogen C ommittee workshop held on O ctober 20-22, 2008. T hese participants included: the following individuals from the U.S. EP A - Mr . Richard Batiuk, D r . Jana C ompton, D r . Robin D ennis, Ms. Katie F lahive, D r . Richard H aeuber , Ms. C hris Lewiciki, D r . James Liebman, Ms. Sally Shaver, Ms. Roberta P arry , Mr. Randy Waite, Mr. John Wilson, Mr. T om Wirth, and Mr . Marcus Zobrist; the following individuals from the US D A - D r

. D an Jaynes, D r . Ray Knighton, and Mr. Richard Swenson; and the following individuals from other organizations - D r . Jan Willem E risman, E ner gy Research C entre of the N etherlands; Mr . Albert E ttinger
, E nvironmental Law and P olicy C enter; Mr . John H ardin, John H ardin & Son; D r . Shelie Miller, University of Michigan; D r

. C atherine O ’ C onnor , Metropolitan Water Reclamation D istrict of Greater C hicago; Mr . Robin O ’Malley
, T he H einz C enter; Mr . Richard P oirot, V ermont Agency of N atural Resources; Mr . John Quinn, C onstellation E ner gy; Mr. P at Rice, N ebraska D epartment of E nvironmental Quality; D r . Joe Rudek, E nvironmental D efense F und; D r

. T ed Russell, Georgia I nstitute of T echnology; Ms. Mindy Selman, W orld Resources institute; D r

. T homas Simpson, University of Maryland; D r . Richard Smith, U.S. Geological Survey; D r

. C lif ford Snyder, I nternational P lant N utrition I nstitute; D r . Robert Summers, Maryland D epartment of the E nvironment; D r

. D avid Whitall, N ational O ceanic and Atmospheric Administration; Ms. Marcia Willhite, I llinois E nvironmental P rotection Agency; and Mr . James Wurtz, Railway E quipment C ompany

. I n addition, the C ommittee thanks the experts who provided independent review of the report in draft form: D r

. John D ay , Louisiana State University; D r

. E lisabeth H olland, N ational C enter for Atmospheric Research; D r . Gregory Mc I saac, University of I llinois; D r . Jerry Melillo, Marine Biological Laboratory, Woods H ole, Massachusetts; D r . Gyles Randall, University of Minnesota; D r . James Schauer, University of Wisconsin-Madison; and D r . Stuart Weiss, C reekside C enter for E arth O bservations. T he reviewers provided many constructive comments and suggestions, but were not, however , asked to endorse the conclusions or recommendations in the report.viii 

### NOTICE

T his report has been written as part of the activities of the EP A Science Advisory Board (SAB), a public advisory group providing extramural scientific information and advice to the Administrator and other officials of the E nv ironmental P ro tection Agency. T he SAB is structured to provide balanced, expert assessment of scientific matters related to problems facing the Agency. T hi s 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 E nv ironmental P ro tection Agency, nor of other agencies in the E xe cutive Branch of the F ed eral government, nor does mention of trade names of commercial products constitute a recommendation for use. Reports of the SAB are posted on the EP A 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 Acr onyms and Abbreviations** 

**. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . xvi**  **Executive Summary**

**. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .**  **ES-1**

**1**

**. Intr**  **oduction**

**. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .**  **3**

1.1. Overvie w 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 Objecti ves

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 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 En vironment

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13

2.2.1. Nr F ormation and Losses to the En vironment fr om F ossil Fuel Comb ustion

. . . . . . . . . . . . . . . . . . . . . . . 14

2.2.2. Nr Inputs and Losses to the En vironment fr om Cr op Agriculture

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15

2.2.3. Nr Inputs and Losses fr om Animal Agriculture

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 25

2.2.4. Nr Inputs to Residential and Recr eational Turf Systems

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 29

2.3. Nr T ransfer and Transformations in and Between Environmental Systems

. . . . . . . . . . . . . . . . . . . . . . . . . . . 31

2.3.1. Input and Tr ansfers of Nr in the United States

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 31

2.3.2. Storage of Nr Within Terr estrial En vironmental Systems

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 35

2.3.3. Areas of Uncertainty in Nr Tr ansfer and Tr ansformation

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 38
**3**

**. Impacts of Nr on**  **Aquatic, Atmospheric, and Terrestrial Ecosystems**

**. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .**  **41**

3.1. Impacts on Drinking W ater, 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 T errestrial Ecosystems

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 42

**4**

**. Metrics and Curr**  **ent 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 Re gulation 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. Tradeof fs of Nr Impacts in Risk Reduction Strategies

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 53

4.9. Interactions of the N Cascade and Climate

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 56 

# Table of Contentsx **5**

**. Integrated Risk Reduction Strategies for Nr** 

**. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .**  **59**

5.1. Importance of Integrated Risk Reduction Strate gies

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 59

5.2. Control Strategies for Nr

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 59

5.3. Management Strategies for Nr in the En vironment

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 59

5.3.1. Command-and-control

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 59

5.3.2. Direct Allocation of F ederal Funds for Conservation Pr ograms

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 60

5.3.3. Market Based Instruments for P ollution Contr ol

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 61

5.3.4. Biophysical and Tec hnical Contr ols (contr ol points) on Tr ansfer and Tr ansformations of Nr in and Between En vironmental Systems

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . 62
**6**

**. SAB Recommendations f**  **or Nr Data Collection, Risk Management, and Research**

**. . . . . . . . . . . . . . . . . . . .**  **73**

6.1. Need for Comprehensiv e 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 F our 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 EP**  **A Mobile Source Rules to Control NO x**  **. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 109**  **Appendix G: Impacts of Reacti ve 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 reactiv e nitrogen (Nr) introduced into the United States in 2002 (Tg N/yr)

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

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

. . . . . . . . . . . . . . . . 11 Figure 3: U.S. NO x emission trends, 1970-2006

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14 Figure 4: Percent reductions in NO x emissions, 1990-2002, from different sources (off-road, on-road, power generation, etc.)

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14 Figure 5: Mobile source NO x 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 liv estock 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. inv entory 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: Relativ e 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) relativ e 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 W arming Potential (GWP) and eutrophication impact categories for various bioproducts

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 56 Figure 20: Probability of giv en discharge level for nitrate in the watersheds of eastern Iowa

. . . . . . . . . . . . . . . . . . . 57 Figure 21: Relativ e nitrogen discharge (lbs/day) from 79 POTWs

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 64 Figure 22: T rading ratios for municipalities in Connecticut

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 65 Figure 23: The lik ely impact of research investment in increasing N fertilizer use efficiency

. . . . . . . . . . . . . . . . . . . 68 Figure A-1: Percent change in relativ e contribution of oxidized (NO 3 - ) and reduced (NH 4 + ) nitrogen wet deposition from 1994 to 2006

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 89 Figure A-2: Trend in reported wet deposition of NH 4 + and NO 3 - for the 48 contiguous states

. . . . . . . . . . . . . . . . . . 90 Figure A-3: Annual NH 4 + , NO 3 - , and total inorganic N deposition for the year 2007 showing spatial patterns of deposition

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 90 Figure A-4: CMAQ annual a verage (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 dischar ge, and industrial dischargers in the Illinois River Watershed

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

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 100 Figure C-3: Potential land av ailability in the 100-year flood zone for nutrient farming in each sub-watershed in the Illinois River Watershed

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 100 Figure C-4: Spring av ailable total nitrogen load by sub-watershed

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 100xii Figure C-5: Spring marginal cost (price) by w atershed

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

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 101 Figure E-1: T otal 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

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 113xiii 

# List of Tables

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

. . . . . . . . . . . . . ES-2

Table 1: Nr flux es for the United States, Tg N in 2002

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12

Table 2: Examples of multiple sources from states with high NO x emissions based on 2001 data; and tons of NO x as NO 2
. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 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: N 2 O emissions in the United States, 2002

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23

Table 6: Li vestock 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: F ate 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: Mar ginal abatement cost per tonne of Nr by source

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 47 Table 13: Federal primary ambient air quality standards that in volve 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 mark et-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 NH 3 emissions from livestock manure in the United States

. . . . . . . . 78

Table A-1: Annual wet deposition of reduced (NH 4 + ), oxidized (NO 3 - ), 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 reactiv e 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 mark et parameters under three trading scenarios

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 101

Table D-1: Summary of the effects of e xcess 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 e xcess 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 e xcess 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

. . . . 118xiv 

# List of Boxes

Box 1: Hypoxia in the Gulf of Mexico

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 35 Box 2: Economic impact and metrics for Chesapeake Bay and its w atershed

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 46 Box 3: The impact of climate change on agricultural dischar ge of reactive nitrogen

. . . . . . . . . . . . . . . . . . . . . . . . . 57

Box 4: W ater quality trading to meet the Long Island Sound wasteload allocation in Connecticut

. . . . . . . . . . . . . . . 63

Box G-1: The national criterion for ammonia in fresh w ater

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 114 Box G-2: Long Island Sound total maximum daily load: focus on reactiv e nitrogen

. . . . . . . . . . . . . . . . . . . . . . . . 119xv 

# List of Chemical Abbreviations

C Carbon

CFC Chlorofluorocarbon

DIN Dissolved inor ganic nitrogen

DO Dissolved oxygen

Fe Iron

H Hydrogen

HNO 3 Nitric acid

HONO Nitrous acid

N Nitrogen

N 2 Diatomic (molecular) nitrogen

N 2 O Nitrous oxide,

N 2 O 5 Dinitrogen pentoxide (nitric acid anhydride)

NH 3 Ammonia

NH 4 + Ammonium

NH x NH 3 NH 4 +

NO Nitric oxide

NO 2 Nitrogen dioxide

NO 3 - Nitrate ion

NO 3 Nitrate radical

N org Org anic nitrogen

NO x Nitrogen oxides (NO + NO 2 )

NO y Total reacti ve oxidized nitrogen (NO, NO 2 , NO 3 , 2xN 2 O 5 , HONO, HNO 3 , NO 3 - , PAN and other organo-nitrates, RONO 2 )

Nr Reactiv e nitrogen

O 2 Oxygen

OH Hydroxyl radical

P Phosphorus

PAN Peroxy acetyl nitrate

PM Particulate matter

PM 2.5 Particulate matter less than 2.5 microns in diameter

PM 10 Particulate matter less than 10 microns in diameter

R ONO 2 Org anic nitrates

Si Silicon

SO 2 Sulfur dioxide

SO 4 2- Sulfate

T

AN Total ammonical nitrogenxvi 

# List of Acronyms and Abbreviations

AAPFCO Association of American Plant F ood Control Officials

AARA American Reinv estment and Recovery Act

AIRMON Atmospheric and Integrated Research Monitoring Netw ork

AOB Ammonia oxidizing bacteria

BL Boundary layer

BMP Best management practice

BNF Biological nitrogen fixation

BNR Biological nutrient (or nitrogen) remov al

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 Cultiv ation-induced biological nitrogen fixation

CCC Criterion continuous concentration

CFC Chlorofluorocarbon

CFR Code of Federal Regulations

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

CLAD Critical Loads Ad-Hoc Committee

CMA Q Community multiscale air quality

CMC Criterion maximum concentration

CRP Conservation Reserv e Program

CSO Combined sewer o verflow

CTM Chemical T ransport Models

CWA Clean W ater Act

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

DOE U.S. Department of Energy

DO T U.S. Department of T ransportation

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 Incenti ves Program

EU European Union

FA O Food and Agricultural Organization of the United Nations

FAOSTAT Food and Agricultural Organization Statistical Database

FGR Flue-gas recirculationxvii ha Hectare

GHG Greenhouse gas

GPS Geographic Positioning System

HAB Harmful algal bloom

IPCC Intergo vernmental 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 lev el

MCLG Maximum contaminant lev el goal

mg Milligrams

MGD Million gallons per day

MJ Meg ajoule (one million joules)

Mmt Million metric tons

MT metric tons

MOM Mississippi-Ohio-Missouri

MRB Mississippi Riv er Basin

MS4 Municipal separate storm sewer system

N

AAQS National Ambient Air Quality Standards

NADP National Atmospheric Deposition Program

N

ASS National Agricultural Statistics Service Information

NCA National Coastal Assessment

NCE Nitrogen Credit Exchange

NCCR National Coastal Condition Report

NEEA National Estuarine Eutrophication Assessment

NESCA UM 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 Inv entory
NTN National T rends Networkxviii 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 org anic matter mineralization

OTAG Ozone T ransport Assessment Group

OTC Ozone T ransport Commission

PE Physiological efficienc y (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 abov e ground crop biomass)

PFP Partial f actor productivity

POTW Publicly owned treatment w orks

PSD Prev ention of significant deterioration

RE Recov ery efficiency (kg N uptake per kg N applied)

SAV Submerged aquatic v egetation

SNCR Selectiv e non-catalytic reduction

SCR Selectiv e catalytic reduction

SIP State Implementation Plan

SOM Soil org anic matter

SPARROW Spatially Referenced Regressions on Watershed Attributes Model

STP Sew age treatment plant

SW Storm water

SW AT Storm W ater Assessment Tool

SWPPP Stormwater Pollution Pre vention Plan

T Temperature Tg Teragram (million metric tons or 10 12 grams)

TMDL Total maximum daily load

TN Total nitrogen

UFTRS Uniform Fertilizer T onnage Reporting System

UNECE United Nations Economic Commission for Europe

USDA U.S. Department of Agriculture

USGS U.S. Geological Surve y
USEPA United States Environmental Protection Agency

WHO World Health Or ganization

WLA Wasteload allocation

WPCA Water pollution control authorities

WRI World Resources Institute

WRP Wetland Reserv e Program

WSA Wadeable Stream AssessmentES-1 

### Introduction
N itrogen ( 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 ( N r), in contrast to non-reactive gaseous N 2 , 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 or ganism’s diet, it can’t survive. I ronically
, 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

. T his is certainly the situation with food production. Without the creation of N fertilizer by an industrial process (the H aber -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 C hapter 2 of this report, increased anthropogenic input of N r to the environment has contributed to lar ge 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. H owever , a major consequence of this nearly inexhaustible supply is that most N used in food production, and all of the new N r produced by fossil fuel combustion, is lost to the environment where it circulates through the earth’ s atmosphere, hydrosphere, geosphere, and biosphere. D uring this circulation, N r contributes to a wide variety of consequences, which are magnified with time as N r moves through the environment.


### Impacts of reactive nitrogen on human health and the environment Anthropogenic creation of N r 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. E ssentially all of the N r created by human activities, however , is released to the environment, often with unintended negative consequences. As summarized in T able E S-1, it contributes to a number of adverse public health and environmental ef fects, 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 N r imbalances, greenhouse gas emissions and subsequent climate change, and stratospheric ozone depletion.

I n light of the magnitude of the human alteration of the nitrogen cycle, and the resulting negative consequences on humans and ecosystems, the N at ional Academy of E ng ineering has identified management of the nitrogen as one of the “grand challenges” facing this country. 2

N r 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). I n addition, there are indirect declines in human health because the negative impacts on ecosystems will diminish the services that those ecosystems provide people. T he ef fects 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
T he nitrogen cascade has three dimensions: biogeochemical, alterations in the environment, and human and ecosystem consequences.

T he “biogeochemical” dimension of the nitrogen cascade involves: N r creation from N 2 as a consequence of chemical, food, and energy production; N r use in food and chemical production; N r losses to the environment; changes in N r species residence times in environmental reservoirs; N r transfers among reservoirs; and N r conversion back to N 2 . Alterations to the environment then result from increased N r levels in the environment. T hese 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. T he overarching question is, how do we protect and sustain ecosystems that provide multiple benefits to society while also providing the interconnected material, food and ener gy required by society? 

# Executive Summary

1 Reactive nitrogen ( N r) includes inor ganic chemically reduced forms of N ( NH x ) [e.g., ammonia ( NH 3 ) and ammonium ion ( NH 4 + )], inorganic

chemically oxidized forms of N [e.g., nitrogen oxides ( NO x ), nitric acid ( HNO 3 ), nitrous oxide ( N 2 O ), N 2 O 5 , HONO , peroxy acetyl compounds such as peroxyacytyl nitrate ( P A

N ), and nitrate ion ( NO 3 - )], as well as organic compounds (e.g., urea, amines, amino acids, and proteins).

2 N ational Academy of E ngineering Grand C hallenges (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 NO y and NH x 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 speciesrichness 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, HNO 3 , NH 3

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 nutrientloading, 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 PM 2.5 , O 3 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). NO y and NH x from fossil fuels and

agriculture Mokdad et al., 2004; Ezzati et al., 2004.

Total damage to public health and

environment NO x 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 NH x 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 environmentES-3 N r inputs to the nation and the world have been increasing, largely due to human activities associated with food production and fossil fuel combustion. D espite the obvious benefits of a plentiful supply of food and

ener gy, the adverse consequences associated with the accumulation of N r in the environment are lar ge, with implications for human health and the environment.

T he greater the inputs of N r to the landscape, the greater the potential for negative ef fects caused by greenhouse gas (G H G) production, ground level ozone, acid deposition, and N r 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.

T he growing nature of the N r problem, and the adverse and intertwined consequences associated with N r 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 N r removal processes, and protect humans and natural resources.

T he concept of the nitrogen cascade highlights that

once a new N r molecule is created, it can, in sequence, travel throughout the environment contributing to major environmental problems (Galloway et al., 2003). T he adaptation of the cascade in F igure E S-1 was developed by the SAB I ntegrated N itrogen C ommittee ( INC ) to provide a context for considering nitrogen-related issues and ecosystem ef fects in the U.S. T o consider the cascading effects of N r in the U.S., we examined the various atmospheric, terrestrial, and aquatic environmental systems where N r is stored, and the magnitudes of the various flows of N to, from, and within them. T he nitrogen cascade concept implies the cycling of N r among these systems. T he process of denitrification is the only mechanism by which N r is converted to chemically inert N 2 , “closing” the continuous cycle ( F igure E S-1 shows only flows of reactive nitrogen, not N 2 ). D enitrification can occur in any of the indicated reservoirs except the atmosphere.

T he “new” N box in the N itrogen C ascade depicts the two primary anthropogenic sources by which N r originates – ener gy production and food production – and where N r from these sources enters ecosystems. E ner gy production includes both fossil fuel and biofuel combustion. F ood production includes N fertilizer produced in the U.S., cultivation-induced biological N ( C -B NF ) 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 NO y , NH x , N org

NO y , NH x , N org stratosphere

troposphere ozone 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 NO y , NH x , N org

N 2 O N 2 O NO x NH 3 N org biogeochemical cycling

ecosystem productivity

acidification

eutrophication NO x

NH xES-4 T he atmospheric system box in the F igure E S-1 indicates that tropospheric concentrations of both ozone and particulate matter are increased due to emissions of nitrogen oxides 3 ( NO x ) to the atmosphere. T he ovals illustrate that the increase in N 2 O concentrations, in turn, contribute to the greenhouse effect in the troposphere and to ozone depletion in the stratosphere. E xcept for N 2 O
, there is limited N r storage in the atmosphere. Losses of N r from the atmospheric system include total oxidized nitrogen 4 ( NO y ), reduced nitrogen 5 ( NH x ), and organic nitrogen ( N or g) deposition to terrestrial and aquatic ecosystems of the earth’s surface. T here is little potential for conversion of N r to N 2 via denitrification in air. H owever , once airborne deposition of N r occurs it will be subject to denitrification pathways via soil and water

. T he terrestrial system box in the F igure E S-1 depicts that N r 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 N r as food to populated regions, from which there are N r losses to the environment (e.g., sewage, landfills). T he ovals showing ecosystem productivity and biogeochemical cycling reflect that N r 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

. T here is ample opportunity for N r storage in both biomass and soils. Losses of N r from this system occur by leaching and runoff of NO y , NH x and N org to aquatic ecosystems and by emissions to the atmospheric system as NO x , NH 3 , N or g, and N 2 O

. T here is potential for conversion of N r to N 2 via denitrification in the terrestrial system.

T he aquatic system box in the F igure E S-1 shows that N r is introduced via leaching and runof f from terrestrial ecosystems and via deposition from atmospheric ecosystems. C onnected with the hydrological cycle, there are N r fluxes downstream with ultimate transport to coastal systems. Within the aquatic system, the ovals highlight two significant impacts of waterborne N r acidification of freshwaters and eutrophication of fresh and coastal waters. E xcept for N r accumulation in groundwater reservoirs, there is limited N r storage within the hydrosphere. Losses of N r from the aquatic system are primarily via N 2 O emissions to the atmospheric system. T here is a very lar ge potential for conversion of N r to N 2 via denitrification in water and wetlands.

NO y , NH x and N 2 O are all components of N r, but a fundamental difference is that the NO y and NH x 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
, N 2 O 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
I n 2002, humans introduced 29 teragrams ( T g) of newly formed reactive N into the U.S. through H aber

Bosch process production of fertilizers and industrial N r , cultivation-induced biological nitrogen fixation (i.e., conversion of N 2 to NH 3 by microorganisms associated with some cultivated crops, for example, legumes), and fossil fuel combustion ( F igure E S-2). By definition, prior to human presence in the U.S., there was no introduced anthropogenic N r

. P rior to 1900, no H aber -Bosch N r was introduced, fossil fuel combustion introduced very small amounts relative to today , and cultivation-induced biological nitrogen fixation created approximately 2 T g N

. T hus, between 1900 and 2002, the amount of N r introduced to the U.S. has increased by approximately 10-fold.


### Nitrogen inputs to the United States
T he EP A Science Advisory Board (SAB) I ntegrated N itrogen C ommittee (“ C ommittee”) evaluated nitrogen inputs to the U.S. in 2002. At the global scale, human activities produced approximately twice as much N r as did natural processes. I n the U.S., however , the amount of N r produced by human activities was approximately five times lar ger than natural processes. As shown in F igure E S-2, natural ecosystems in the U.S. introduce about 6.4 T g of N r as N per year ( T g N /yr). I n contrast, human activities introduce about 28.5 T g N /yr

. C hapter 2 of this report discusses sources, transfer
, and transformation of N r . Supporting references for the information presented on this topic are presented in C hapter 2. T he lar gest single source of N r in the U.S. is the H aber -Bosch process, which introduces about 15.2 T g N /yr: 9.4 T g N /yr from domestic N r production and 5.8 T g N /yr from imports of N r in fertilizers. T he 15.2 T g N /yr of anthropogenic N r is used in three ways: 9.9 T g N /yr is used to produce agricultural crops; 1.1 T g N / yr is applied to turf grasses; and 4.2 T g N /yr is used by industry for production of nylon, refrigerants, explosives and other commercial products.

T he second lar gest source of N r introduced into the U.S. is enhancement of biological nitrogen fixation (B NF ) 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. T hese N r fixing crops introduce about 7.7 T g N /yr . A

3 NO x (oxides of nitrogen) includes NO + NO 2 4 NO y (total oxidized nitrogen) includes NO
, NO 2 , NO 3 , N 2 O 5 , HONO
, HNO 3 , NO 3 –, P A

N and other or gano-nitrates, R ONO 2 5 NH x (reduced nitrogen) includes NH 3 + NH 4ES-5 small amount of additional N r is also imported in grain and meat products; in 2002 this source of added N r was approximately 0.2 T g N /yr (not shown in F igure E S-2).

F ossil fuel combustion is the third lar gest source of

new N r

. I t introduces approximately 5.7 T g N /yr into the environment (almost entirely as NO x ), that is, 3.8 T g N /yr from transportation sources and 1.9 T g N /yr from stationary sources such as electric utilities, industrial boilers and from certain industrial processes.

I n summary, agriculture and domestic use of fertilizers to produce food, feed, and fiber (including bioenergy and B NF ) and combustion of fossil fuels are the largest sources of new N r released into the environment in the U.S. T he percentage distribution of N r released to the U.S. environment from human activities in 2002 was: about 65% from agricultural sources (including B NF a nd turf production), about 20% from fossil fuel sources, and about 15% from industrial sources ( F ig ure E S- 2).


### Distribution of reactive nitrogen through the environment
O nce introduced into the U.S., N r compounds are distributed via the atmosphere, hydrosphere, geosphere, biosphere, and commerce. D istribution in the atmosphere begins with NO x , NH 3 , and N 2 O

. NO x and NH 3 (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. D ue to its long lifetime (approximately 100 years) in the atmosphere, N 2 O 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: N umerical units: teragram of reactive nitrogen ( N r) per year ( T g N /yr)

N atural B NF : biological nitrogen fixation in natural grasslands, rangelands, and forests,

F ossil F uel- T ransportation: combustion in vehicles, trains, airplanes, ships and of f-road construction equipment.

F ossil F uel-Stationary: combustion of fossil fuels in power plants and industrial boilers. Agriculture-cultivation B NF : agricultural augmentation of biological nitrogen fixation – for example by planting of nitrogen fixing legumes. Agriculture- H aber Bosch N fertilizer: agricultural (including turf production) use of synthetic nitrogen fertilizers produced by the H aber Bosch process for converting gaseous N 2 to N r.
I ndustry- H aber Bosch N : I ndustrial sources of N r produced by the H aber -Bosch process.

F igure E S-2 documents only the introduction of new N r in the United States, and not the transfers of existing N r among systems (e.g., N r in manure).ES-6 and is also dispersed throughout the global atmosphere. All ecosystems in the conterminous U.S. receive anthropogenic N r from the atmosphere and for many ecosystems it is their primary , albeit unintended, source of N r

. O nce deposited, N r can be stored in soils and biomass and widely distributed via the stream-river continuum to inland and coastal waters. Some of the N r is converted to N 2 O or denitrified to N 2 , primarily in aquatic ecosystems, including wetlands. C ommerce is a major mechanism that transfers N r from one place to another in the U.S.; most of the N r that is used to produce food (e.g., fertilizer) and in food products crosses state boundaries via roads, railroads and the air

. P ut ting values to this distribution, of the 6.3 T g N /
yr of U.S. NO x emissions, 2.7 T g N /y r are deposited back onto the land and surface waters of the U.S. T hu s, by difference we estimate that as much as 3.6 T g N /
yr of the U.S. NO x emissions are advected out of the U.S. via the atmosphere. Similarly, of the 3.1 T g N /y r of NH 3 that are emitted into the U.S. atmosphere each year, about 2.1 T g N /y r are deposited onto the land and surface waters of the U.S., and about 1 T g N /y r is advected out of the U.S. via the atmosphere. E mi ssions of N 2 O discharge about 0.8 T g N /y r into the global atmosphere. I n sum, 5.4 T g N a re 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 N r to the U.S. coastal zone account for 4.8 T g N /yr , while export of N -containing commodities (e.g., grain) removes another 4.3 T g N / yr from the U.S.. Altogether, along with 5.4 T g N /yr of atmospheric advection, these total N r outputs out of the U.S. continental environment add up to about 14 T g N /yr , leaving about 21 T g N /yr unaccounted for

. O f this amount, we estimate that 5 T g N /yr are stored in soils, vegetation, and groundwater and, by dif ference, we estimate that about 16 T g N /yr are denitrified to N 2
. D enitrification, a process that microbially converts N r to N 2 (as well as forming some N 2 O ) 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. T his process can be a major N r sink in river basins. T here 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 NH x . T he C ommittee considered these uncertainties in developing the “ O verarching Recommendations” of this report.


### Current EPA Nr risk management and research programs
T he parts of EP A most directly concerned with managing or conducting research on N r are the O ffi ce of Air and Radiation, the O ffi ce of Water, and the O ffi ce of Research and D ev elopment ( O R D ). O ve r a dozen programs of EP A’ s O ffi ce of Air and Radiation reduce risks from N r. T he se programs and related activities include: N at ional Ambient Air Quality Standards standard setting and implementation; emission standards for industrial stationary sources and area sources; the Acid Rain P ro gram; the C le an Air I nt erstate Rule; and programs that focus on mobile source emissions. P ro grams designed to save energy, such as E ne rgy Star, tend to reduce emissions of N r as well. EP A’ s O ffi ce of Water addresses N r under both the C le an Water Act and the Safe D ri nking Water Act through activities such as: criteria development and standard setting; total maximum daily load ( T M D L) development; N at ional P ol lution D is charge E li mination System ( NPDE S) permits; infrastructure financing through the D ri nking Water and C le an Water State Revolving F un ds; watershed planning; wetlands preservation; and regulation of stormwater and runoff sources that include municipal separate storm sewer systems (MS4), and concentrated animal feeding operations ( C A FO s)

. EP A’ s O ffi ce of Research and D ev elopment aims to conduct leading-edge research and foster the sound use of science and technology in support of the Agency’s mission. T he O ffi ce of Research and D ev elopment is well recognized for providing a scientific basis for the development of the N at ional Ambient Air Quality Standards for NO x and particulate matter ( P M)

. T he O ffi ce of Research and D ev elopment’s E co system Services Research P ro gram 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. T hi s research will support policy and management decisions in EP A’ s O ffi ces of Air and Radiation and Water.

EP A has brought a great variety of risk reduction tools to bear on N r: conventional regulation and enforcement; cap and trade approaches; measurement, monitoring and place-based approaches; control technology development and verification; communication and education; inter governmental and international cooperation; and voluntary approaches. T he variety and breadth of EP A programs addressing N r reflect the ubiquity of N r in the environment, the historical single-medium regulatory approach, and the lack of a “silver bullet” for reducing risks from N r

. 

### Need for an integrated management strategy
T he EP A programs discussed above (and the programs

o f EP A’ s predecessor organizations) have been active in the management of N r through efforts to: decrease or transform N r in sewage; control NO x to decrease photochemical smog and acid rain; control N r inputs to coastal systems; control fine particulates in the atmosphere; and decrease N r leaching and runoff from crop and animal production systems and developedES-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. T hu s there is a need to integrate N m anagement programs, to ensure that efforts to lessen the problems caused by N i n 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 N r

. I ncreasing corn production for ethanol has raised the prospect of increased N r losses (i.e., transfer from fertilized land to water) and degraded water quality

. T he alternative of cellulosic based ethanol does not necessarily mitigate the potential for this negative externality

. H igh yields of cellulosic materials also require N and the “mar ginal” land assumed for such production may be more susceptible to nutrient leakage. Another good example is provided in C hapter 4 of this report (Box 2). T his example considers the water impacts of N r in the C hesapeake Bay and shows that the total reduction of damage from excess N r may rely nearly as much on stricter enforcement of the C lean Air Act as the C lean Water Act. T his challenges the traditional approach to regulation, but it is a consequence of comprehensively examining N r guided by the nitrogen cascade.

T he re can be many unintended consequences associated with a focus on managing one pollutant, even an integrated focus on various forms of N

. F or example, as further discussed in C ha pter 4 and Appendix G of this report, numerous lakes, reservoirs, rivers, estuaries (e.g., the Gulf of Mexico), and fjords worldwide exhibit N a nd phosphorus ( P ) co-limitation, either simultaneously or in seasonally-shifting patterns. T he refore, strategies are needed to reduce both P a nd N i nputs, 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. EP A SAB, 2007). An integrated strategy should take this into consideration.


### Objectives of the SAB Integrated Nitrogen Committee study
T he EP A Science Advisory Board formed the I ntegrated N itrogen C ommittee to assist EP A in its understanding and management of nitrogen-related air , land, and water pollution issues. I n this report, the C ommittee has provided findings and recommendations addressing the following objectives. Assessment of the challenges and costs to EP A 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. T

o address this objective, the C ommittee used the nitrogen cascade framework to determine the major sources of newly created N r in the U.S. ( F igure E S-1). T he flows of N r 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 N r is lost to the environment. T he same process was employed for fossil fuel ener gy production but, since all the N r formed and released during ener gy production is lost to the environment, the C ommittee identified the important ener gy producing sectors that contribute to N r emissions.

T he C ommittee next examined the fate of the N r lost to the environment, estimated the amount stored in dif ferent systems (e.g., forest soils) and tracked N r as it is transferred from one environmental system (e.g., the atmosphere) to another (e.g., terrestrial and aquatic ecosystems).

S ource and fate analyses set the stage for identifying the environmental and human health problems N r presents, and the links among them. Using the nitrogen cascade, the C om mittee identified the impacts N r has on people and ecosystem functions as it moves through each system. T he C om mittee also addressed the alternative metrics that could be used, including the number of tons of specific forms of N r, 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 ofecosystem services (e.g., loss of biodiversity), and tradeoffs among N r impacts. 2. Evaluate the contribution an integrated nitrogen management strategy 6 could make to environmental protection. An integrated management strategy should take into account the contributions of all N r sources, and all chemical species of N r that adversely impact both human health and environmental systems. F urther , an integrated strategy should ensure that solving one problem related
to N r does not exacerbate another problem or diminish ecosystem services that support societal demands. I n short, the strategy should seek to achieve desirable benefits of N r , while limiting adverse effects.ES-8 T o address this challenge, the C ommittee identified several actions that could be taken to better manage N r in one environmental system and avoid unintended consequences in another

. E xamples of “integrative” management actions that could be taken are highlighted in that discussion. 3. Identify additional risk management options for EP A’s consideration. As further discussed below, the C om mittee has identified four major goals for management actions that collectively have the potential to decrease N r losses to the environment by about 25 percent. D ec reasing N r emissions by these actions will result in further decreases in N r- related impacts throughout the nitrogen cascade. T he C om mittee 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. T he se 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. T hu s, the last sections of this report focus on a better understanding of N r 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. I n this report, the C ommittee has provided numerous recommendations for additional N r research to support risk reduction activities. T hese research recommendations are discussed in various chapters of the report and are consolidated in the summary of findings and recommendations presented in C hapter 6.


### Major Findings and Recommendations
T hroughout the report there are boxes containing summary statements labeled “ F indings.” Attached to these findings are one or more specific “Recommendations” for actions that could be taken by EP A or other management authorities. I n each case, the intent is to provide the scientific foundation regarding a specific N r -relevant environmental issue and one or more recommendations by which EP A acting alone or in cooperation with other

or ganizations could use currently available technology to decrease the amount of N r lost to the U.S. environment. T he findings and recommendations are consolidated in C hapter 6 of this report.

O verarching recommendations O pt imizing the benefits of N r, and minimizing its impacts, will require an integrated nitrogen management strategy that involves action not only on the part of EP A, b ut also coordination with other federal agencies, the states, the private sector, universities, and the public, supported by a strong public outreach program. T he refore the C om mittee has also provided four overarching recommendations to assist EP A in its understanding and management of nitrogen-related air, land, and water pollution issues: Overarching Recommendation 1: T he C om mittee recommends an integrated approach to the management of N r. T hi s approach draws upon a combination of implementation mechanisms. E ac h mechanism must be appropriate to the nature of the problem at hand, be supported by critical research on decreasing the risks of excess N r, 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. T hi s is why understanding the nature and dynamics of the N c ascade is critically important. Overarching Recommendation 2: T he framing of the reactive nitrogen cascade provides a means for tracking nitrogen as it changes form and passes through multiple ecosystems and media. T hi s 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 N r. I t 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. T hi s in turn will require a monitoring system that will provide feedback on the effectiveness of specific actions taken to lower fluxes and concentrations of N r.
O verarching Recommendation 3: An intraAgency N r management task force within EP A is recommended to build on existing N r research and management capabilities within the Agency. T hi s task force should be aimed at increasing scientific understanding of: (1) N r impacts on terrestrial and aquatic ecosystems, human health, and climate; (2) N r- relevant monitoring requirements; and (3) the most efficient and cost-effective means by which to decrease various adverse impacts of N r loads as they cascade through the environment. Overarching Recommendation 4: Successful N r

ma nagement will require changes in the way EP A i nteracts with other agencies. C oo rdinated federal programs could better address N r concerns and help ensure clear responsibilities for monitoring, modeling, researching and managing N r in the environment. T hu s, the C om mittee recommends that EP A convene an inter-agency N r management task force. I t is recommended that the members of this inter-agency task force include at least the following federal agencies: U.S. D ep artment of Agriculture,ES-9 U.S. D epartment of E ne rgy, U.S. D ep artment of H ou sing and Urban D ev elopment, U.S. D ep artment of T ra nsportation, N at ional O ce anic and Atmospheric Administration, U.S. Geological Survey, U.S. F or est Service, and F ed eral E me rgency Management Agency. T he EP A O ff ice of I nt ernational and T ri bal Affairs should work closely with the D ep artment of State to ensure that EP A is aware of international efforts to control N r 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 N at ional Research C ou ncil’s recent reports on the Mississippi Basin ( N R C
, 2008b, 2009). T he se intra- and inter-agency N r management task forces should take a systems approach to research, monitoring, and evaluation to inform public policy related to N r management, and implement a systems approach to N r management, as recommended by the C om mittee.


### Summary of specific recommendations by study objective
T he C ommittee’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. T he C om mittee finds that uncertainty associated with rapid expansion of biofuels, losses of N r 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 N r consistently and accurately. Addressing this need will decrease uncertainty in the understanding of the fate of N r that is introduced into the environment and lead to a better understanding of the impacts of excess N r on the health of people and ecosystems. T hi s should be accomplished through a coordinated effort among cognizant federal and state agencies, and universities.

I n addition, the C om mittee recommends that EP A

r outinely and consistently account for the presence of N r 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 N at ional Atmospheric D ep osition P ro gram summary reports). T he C om mittee 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 N r creation, transport and transformation, impacts, and management. 2. Evaluate the contribution an integrated nitrogen management strategy could make to environmental protection. T he C om mittee finds that effective management of N r in the environment must recognize the existence of tradeoffs across a number of impact categories involving the cycling of nitrogen and other elements. I n addition, an integrated multi-media approach to monitoring N r is needed.

I n that regard, the C ommittee recommends that:

1. EP A should develop a uniform assessment and management framework that considers the ef fects of N r loading over a range of scales reflecting ecosystem, watershed, and regional levels. T he framework should include all inputs related to atmospheric and riverine delivery of N r to estuaries, their comprehensive ef fects on marine eutrophication dynamics, and their potential for management.

2. EP A should examine the full range of traditional and ecosystem response categories, including economic and ecosystem services, as a basis for expressing N r 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. T he C ommittee finds that a number of risk management actions should be considered to reduce N r loading and transfer to the environment. T hese include farm-level improvements in manure management, actions to reduce atmospheric emissions of N r , and interventions to control N r in water management programs. As an example, the C ommittee recommends that EP A should reexamine the criteria pollutant “oxides of nitrogen” and the indicator species NO 2 and consider supplementing this with NH x and NO y as indicators of chemically reactive nitrogen ( N r without N 2 O ).

4. Make recommendations to EPA concerning improvements in nitrogen research to support risk reduction. T he C ommittee finds that research is needed in a number of areas to support N r risk reduction activities. T hese areas include research to advance the understanding of: the quantity and fate of N r applied to major crops; how to accelerate crop yields while increasing N fertilizer uptake efficiency; agricultural emissions of forms of N r; atmospheric deposition of N r; and the potential for amplification of N r -related climate impacts.


### Four recommended management actions C onsistent with the overarching and specific

r ecommendations noted above, the C om mittee 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 N r in the environment.ES-10 1. T he C ommittee estimates that if EP A were to expand its NO x 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. E missions from many point sources are controlled with low- NO x burners or NO x reduction. Such equipment should also be installed on industrial boilers and the remaining, uncontrolled power plants. NO x 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. T he C ommittee estimates that excess flows of N r 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. T his would include activities such as using lar ge-scale wetland creation and restoration to provide needed ecosystem services of N r retention and conversion as well as matching cropping systems and intensity of N r use to land characteristics. I mproved tile-drainage systems and riparian buf fers on cropland, and implementing stormwater and non-point source management practices (e.g., EP A permitting and funding programs) are important components. I n addition, the C ommittee 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). C rop output can be increased while decreasing total N r by up to 20% of applied artificial N r , amounting to ~ 2.4 Tg N/yr below current amounts of N r additions to the environment. T hese are appropriate actions that could be taken with today’ s available technologies and further progress is possible.

3. T he C ommittee estimates that livestock-derived NH 3

emissions can be decreased by 30% (a decrease of 0.5 Tg N/yr ) by a combination of BM P s and engineered solutions. T his is expected to decrease P M 2.5 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 NH 3 emissions derived from fertilizer applications can be decreased by 20% (decrease by approximately 0.2 Tg N/yr ), through BM P s that focus on improvements related to application rate, timing, and placement.

4. T he C om mittee 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 N r emissions by between 0.5 and 0.8 Tg N/yr . Additional N r management from eligible stormwater and nonpoint sources could be accomplished through increased support.

I mplementing these suggestions will decrease the amount of N r introduced into the United States by about 25%, which will similarly decrease the amount of N r lost to the atmosphere, soils and waters. T he C ommittee 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 

# Reactive Nitrogen in the United States
An An al y s i s of I np u t s
, Fl o ws
, C on s e qu ence s
, a nd M a n ag e m ent O ption s23 

# 1 Introduction

1.1. Overview of the Problem – Impacts of Excess Nr on Human Health and the Environment
N itrogen is an essential nutrient that governs the growth and reproduction of living organisms. Reactive nitrogen ( N r), in contrast to non-reactive gaseous N 2 , includes all biologically active, chemically reactive, and radiatively active nitrogen ( N ) compounds in the atmosphere and biosphere of the E arth. Anthropogenic creation of N r provides essential benefits for humans – first and foremost in meeting human dietary needs. I n 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. H owever , excess releases of N r 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 nitr ogen (Nr) includes inorganic chemically reduced forms of N (NH x ) [e.g., ammonia (NH 3 ) and ammonium ion (NH 4 +)], inorganic chemically oxidized forms of N [e.g., nitrogen oxides (NO x ), nitric acid (HNO 3 ), nitrous oxide (N 2 O), N 2 O 5 , HONO, peroxy acetyl compounds such as peroxyacytyl nitrate (PAN), and nitrate ion (NO 3 -)], as well as organic compounds (e.g., urea, amines, amino acids, and proteins).

T he negative consequences of N r flux in the U.S. environment include increases in photochemical smog and atmospheric particulate matter ( P M 2.5 ), 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 (G H G) 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; H ec k et al., 1984; P ae rl, 1988; MacKenzie and E l- Ashry, 1990; ECOH AB , 1995; Bricker et al., 1999; Rabalais et al., 1999; N R C
, 2000; Mitsch et al., 2001; F en n et al., 2003; E zz ati et al., 2004; Mokad et al., 2004; V er ity et al., 2006; U.S. EP A C le an Air Scientific Advisory C om mittee, 2008; U.S. EP A SAB, 2008; Bobbink et al., 2010). I n light of the magnitude of the human alteration of the nitrogen cycle, and the resulting negative consequences on humans and ecosystems, the N at ional Academy of E ng ineering has identified management of the nitrogen as one of the “grand challenges” facing this country ( N at ional Academy of E ng ineering, 2008).


### 1.2. The Nitrogen Cascade – Nr Loading, Cycling, and Exposure Approximately 78% of the atmosphere is diatomic nitrogen (N 2 ), which is unavailable to most organisms because of the strength of the triple bond that holds the two N a toms together. O ve r evolutionary history, only a limited number of species of bacteria and archaea have evolved the ability to convert N 2 to N r via biological N fi xation. T hu s, even with adaptations to use N e fficiently, many ecosystems of the world are limited by N

. Anthropogenic creation of Nr N itrogen limitation of ecosystem production has driven humans to use increasingly sophisticated and

ener gy-intensive measures to obtain N r to sustain food production and to produce other commodities (e.g., nylon, explosives). I n pre-history , hunters and gatherers harvested food from natural stocks. With the advent of agriculture, local sources of N r were used (soil stocks, crop residue, and manures) to increase productivity of landscapes. I n the nineteenth century , long-range transport of N r to sustain food production increased with the shipment of bird guano from the P acific I slands and nitrates from South America to E urope and other locations. By the beginning of the twentieth century , these sources were not sufficient to sustain the growing global population requirements for food.

T his deficiency led to what has been called one of the world’ s most important discoveries – how to extract N 2 from the atmosphere and convert it to ammonia ( NH 3 ) – called the H aber-Bosch process (Smil, 2001; E risman et al., 2008). Today, this process and cultivationinduced biological N fixation ( C -B NF ) introduce over 140 teragrams ( T g) of N per year (hereafter expressed

as T g N /yr) into the global environment to increase food production (Galloway et al., 2008). Another 23 T g N / yr are introduced by the H aber -Bosch process for the chemical industry, and 25 T g N /yr are introduced via the combustion of fossil fuels (Galloway et al., 2008).

T he total global anthropogenic N r creation rate is ~190 T g N /yr (2005), substantially lar ger than the median of estimates for N r creation by natural terrestrial processes (~100 T g N /yr) (Galloway et al., 2008). T he fact that humans are more ef fective than nature in N r creation means that on average, humans are less reliant on natural sources of N r

. H owever , with global commodity stocks running at a 58-day supply and food prices increasing4 dramatically, the challenge is to increase the nutrient use efficiency of N r in agricultural systems while maintaining or increasing yields (US D A, E RS/W orld Agricultural O utlook Board, July 1 1, 2008. World Agricultural Supply and D emand E stimates).

T here are lar ge regional disparities in N r creation rates on both absolute and per capita bases. T otal N r creation is lar ger in Asia than in any other region. P er capita N r creation is lar gest in N orth America and E urope. H umans also redistribute lar ge amounts of N r among countries or regions of the world through exports of fertilizers, feed grains, and fossil fuels. N evertheless, there are lar ge regions of the world with populations approaching one billion, where there is malnutrition in part due to a lack of available N r to sustain crop production.

T he introduction of N r into most regions of the United States by humans has greatly increased food availability

. H owever , since essentially all the N r created for food production and by fossil fuel combustion is lost to the environment, it has also greatly increased the contribution of N r to a wide variety of environmental problems. Most plants, animals, and microor ganisms are adapted to efficiently use and retain N r . Addition of N r to most ecosystems may first lead to increased uptake, growth, and storage – and hence to increased biomass, including food or fiber production. H owever , further addition of N r in excessive amounts often leads to imbalances in the movement of N r among reservoirs and potential losses 7 to the environment in the form of air emission or water discharges into other ecosystems where N r may disrupt ecosystem functions and have a negative impact on

resources. I n essence, the assimilative capacity of the ecosystem may be insufficient to benefit from increases in N r without disruptive changes.

T hese 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 N r 8 or the “nitrogen cascade” ( F igure 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.

T he nitrogen cascade has three dimensions: biogeochemical, alterations in the environment, and human and ecosystem consequences.

T he “biogeochemical” dimension of the nitrogen cascade involves: N r creation from N 2 as a consequence of chemical, food and energy production; N r use in food and chemical production; N r losses to the environment; changes in N r species residence times in environmental reservoirs; N r transfers among reservoirs; and N r conversion back to N 2 . Alterations to the environment then result from increased N r levels in the environment. T hese 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. T he overarching question is, how do we protect and sustain ecosystems that provide multiple benefits to society while also providing the interconnected material, food and ener gy required by society?

N r inputs to the nation and the world have been increasing, lar gely due to human activities associated with food production and fossil fuel combustion. D espite the obvious benefits of a plentiful supply of food and

ener gy, the adverse consequences associated with the accumulation of N r in the environment are lar ge, with implications for human health and the environment.

T he greater the inputs of N r to the landscape, the greater the potential for negative ef fects, caused by greenhouse gas (G H G) production, ground level ozone, acid deposition, and N r 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.

T he growing nature of the N r problem, and the adverse and intertwined consequences associated with N r 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 N r removal processes and protect humans and natural resources.

T he concept of the nitrogen cascade highlights that

once a new N r molecule is created, it can, in sequence, travel throughout the environment contributing to major environmental problems (Galloway et al., 2003). T he adaptation of the cascade in F igure 1 was developed by the SAB I ntegrated N itrogen C ommittee ( INC ) to provide a context for considering nitrogen-related issues and ecosystem ef fects in the U.S. T o consider the cascading effects of N r in the U.S., we examined the various atmospheric, terrestrial, and aquatic environmental

7 I n the context of this report, “losses” refers to transfers among systems and no t the conversion of N r to N 2 . Whenever N 2 formation is discussed, it is

explicitly stated.

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

atmosphere, N r inputs to grasslands and forests, N r 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). T hresholds are used to determine the amount of excess N r that causes negative ef fects on ecosystem services and functions, and human health. T hresholds vary by metric (e.g., concentration, loading, etc) and depend on the environmental system (e.g., atmosphere, forest). E xamples for specific thresholds are given later in the report in relevant sections.5 systems where N r is stored, and the magnitudes of the various flows of N to, from, and within them. T he nitrogen cascade concept implies the cycling of N r among these systems. T he process of denitrification is the only mechanism by which N r is converted to chemically inert N 2 , “closing” the continuous cycle ( F igure 1 shows only flows of reactive nitrogen, not N 2 ). D enitrification can occur in any of the indicated reservoirs except the atmosphere.

T he “new” N box in F igure 1 depicts the two primary anthropogenic sources by which N r originates – ener gy production and food production – and where N r from these sources enter ecosystems. E ner gy production includes both fossil fuel and biofuel combustion. F ood production includes N fertilizer produced in the U.S., cultivation-induced biological N ( C -B NF ) 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.

T he atmospheric system box in F igure 1 indicates that tropospheric concentrations of both ozone and particulate matter are increased due to emissions of nitrogen oxides 9 ( NO x ) to the atmosphere. T he ovals illustrate that the increase in N 2 O concentrations, in turn, contribute to the greenhouse effect in the troposphere and to ozone depletion in the stratosphere. E xcept for N 2 O , there is limited N r storage in the atmosphere. Losses of N r from the atmospheric system include total oxidized nitrogen 10 ( NO y ), reduced nitrogen 11 ( NH x ), and organic nitrogen ( N org ) deposition to terrestrial and aquatic ecosystems of the earth’s surface. T here is little potential for conversion of N r to N 2 via denitrification in air. H owever, once airborne deposition of N r occurs it will be subject to denitrification pathways via soil and water

. T he terrestrial system box in F ig ure 1 depicts that N r 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 N r as food to populated regions, from which there are N r 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 NO y , NH x , N org

NO y , NH x , N org stratosphere

troposphere ozone 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 NO y , NH x , N org

N 2 O N 2 O NO x NH 3 N org biogeochemical cycling

ecosystem productivity

acidification

eutrophication NO x

NH x

9 NO x (oxides of nitrogen) includes NO + NO 2
10 NO y (total oxidized nitrogen) includes NO
, NO 2 , NO 3 , N 2 O 5 , HONO
, HNO 3 , NO 3 - , P A

N and other or gano-nitrates, R ONO 2
11 NH x (reduced nitrogen) includes NH 3 + NH 46 landfills). T he ovals showing ecosystem productivity and biogeochemical cycling reflect that N r 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. T he re is ample opportunity for N r storage in both biomass and soils. Losses of N r from this system occur by leaching and runoff of NO y , NH x and N org to aquatic ecosystems and by emissions to the atmospheric system as NO x , NH 3 , N org , and N 2 O

. T here is potential for conversion of N r to N 2 via denitrification in the terrestrial system.

T he aquatic system box in F ig ure 1 shows that N r is introduced via leaching and runoff from terrestrial ecosystems and via deposition from atmospheric ecosystems. C on nected with the hydrological cycle, there

are N r fluxes downstream with ultimate transport to coastal systems. Within the aquatic system, the ovals highlight two significant impacts of waterborne N r acidification of freshwaters and eutrophication of fresh and coastal waters. E xc ept for N r accumulation in groundwater reservoirs, there is limited N r storage within the hydrosphere. Losses of N r from the aquatic system are primarily via N 2 O

em issions to the atmospheric system. T he re is a very large potential for conversion of N r to N 2 via denitrification in water and wetlands.

NO y , NH x and N 2 O are all components of N r, but a fundamental difference is that the NO y and NH x 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
, N 2 O 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
EP A activities to manage the risks posed by reactive nitrogen can be linked to the Agency’s broad strategic goals. EP A’ s mission is to protect human health and the environment. I n achieving this mission, EP A is accountable for addressing five goals given in the 2006 – 201 1 EPA Strategic Plan (U.S. EP A, 2006d):

1. C lean air and global climate change

2. C lean and safe water 3. Land preservation and restoration

4. H ealthy communities and ecosystems

5. C ompliance and environmental stewardship

T he Strategic Plan includes targets for reducing risk from N

. EP A’ s Report on the Environment (U.S. EP A, 2008c), provides “data on environmental trends,” to determine whether or not EP A is on track to meet its tar gets and goals. EP A is responsible and accountable for reducing at least some risks from N r

. As previously discussed, the principal mechanisms for N r removal from circulation in the environment are complete denitrification (re-conversion of N r back to nonreactive gaseous N 2 ), and storage in long-term reservoirs (e.g., soils, sediments, and woody biomass). I n some cases, it may be possible to capture N r emissions or discharges and deliver them to food or fiber production areas where there are nitrogen deficiencies. H ow ever, as previously noted, major challenges in the management of the N c ycle are how to decrease creation of N r while still meeting societal needs, promote denitrification of excess N r (w ithout producing N 2 O ), and improve the efficiency of use

a

nd reuse of excess N r in a cost-effective manner. Solving these challenges will result in less N r accumulation.

T he parts of EP A most directly concerned with managing or conducting research on N r are the O ffice of Air and Radiation, the O ffice of Water, and the O ffice of Research and D evelopment. P rograms designed to save

ener gy, such as E ner gy Star, tend to reduce emissions of N r as well. I n over a dozen programs, EP A’ s O ffice of Air and Radiation reduces risks from N r

. T hese programs and related activities include:

n N ational Ambient Air Quality Standards ( N AAQS)

standard setting and implementation;

n E mission standards for industrial stationary sources and

area sources

n Acid Rain P rogram

n C lean Air I nterstate Rule

n P rograms that focus on mobile source emissions

EP A’ s O ffice of Water addresses N r under both the C lean Water Act and the Safe D rinking Water Act through activities such as:

n C riteria development and standard setting

n T otal maximum daily load ( T M D L) development

n N ational P ollution D ischar ge E limination System

( NPDE S) permits

n Watershed planning

n Wetlands preservation

n Regulation of concentrated animal feeding operations

( C A FO s).

EP A’ s O ffice of Research and D evelopment aims to conduct leading-edge research and foster the sound use of science and technology in support of EP A’ s mission. T he O ffice of Research and D evelopment is well recognized for providing a scientific basis for the development of the N ational Ambient Air Quality Standards for NO x and particulate matter ( P M). T he O ffice of Research and D evelopment’ s E cosystem Services Research P rogram has been developed to identify and quantify the positive7 and negative impacts on ecosystem services resulting from changes in nitrogen loadings from major source categories to support policy and management decisions in EP A’ s O ffices of Air and Radiation and Water.

EP A has brought a great variety of risk reduction tools to bear on reactive N :

n C onventional regulation and enforcement

n C ap and trade approaches

n Measurement, monitoring and place-based approaches

n C ontrol technology development and verification

n C ommunication and education

n I nter governmental and international cooperation

n V oluntary approaches

T he variety and breadth of EP A programs addressing N r reflect the ubiquity of N r in the environment, the historical single medium regulatory approach, and the lack of a silver bullet for reducing risks from N r

. Need for an Integrated Nitrogen Management Strategy T he EP A programs discussed above (and the programs of EP A’ s predecessor organizations) have been active in the management of N r through efforts to: decrease the N r

am ount in sewage, control NO x to decrease photochemical smog and acid rain, control N r inputs to coastal systems, control fine particulates in the atmosphere, and decrease N r 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. T hu s, there is a need to integrate N m anagement programs, to ensure that efforts to lessen the problems caused by N i n one area of the environment do not result in unintended problems in other areas. Biofuels feedstock production is a good example of this. I nc reasing corn production for ethanol raised the prospect of increased N r losses and degraded water quality. T he alternative of cellulosic based ethanol does not necessarily mitigate the potential for this negative externality. H ig h yields of cellulosic materials also require N a nd the “marginal” land assumed for such production may be more susceptible to nutrient leakage ( N R C
, 2008a).

I n addition, there can be unintended consequences associated with a focus on one pollutant, even an integrated focus on various forms of nitrogen. F or example, as further discussed in C ha pter 4 and Appendix G of this report, numerous lakes, reservoirs, rivers, estuaries (e.g., the Gulf of Mexico), and fjords worldwide exhibit N a nd P c olimitation, either simultaneously or in seasonally-shifting patterns. T he refore, strategies are needed to reduce both P

an d N i nputs, and not all control practices will be effective for dual nutrient reduction. T he re 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. EP A SAB, 2007) and an integrated strategy should take this into consideration.


### 1.4. SAB Integrated Nitrogen Committee Study Objectives
T he EP A Science Advisory Board has previously
p rovided advice concerning management of nitrogenous compounds as well as integrated environmental decision making. I n 1973, the Science Advisory Board issued a report Nitrogenous Compounds in the Environment (U.S. EP A SAB, 1973). T he report addressed sources and effects of nitrogenous compounds, including those from air emissions, animal wastes, crop agriculture, industrial processes, and solid wastes. T he 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 N r research and control. I n its 2000 report, Toward Integrated Environmental Decision-Making (U.S. EP A SAB, 2000) the SAB articulated a framework for integrated environmental decision-making. I n that report, the SAB noted that the three-phased structure of theframework (problem formulation, analysis and decisionmaking, followed by implementation and evaluation) “belies the complexities involved in putting the concept of integrated decision-making into practice.” T he SAB’s interests in N s cience and integrated environmental protection converged in 2007, when the SAB identified integrated N r esearch and control strategies as an important issue facing the Agency and formed the I nt egrated N it rogen C om mittee (the C om mittee) to conduct this study.

T he C ommittee was char ged by the Science Advisory Board to address the following four objectives:

1. I dentify and analyze, from a scientific perspective, the

problems reactive nitrogen presents in the environment and the links among them;

2. E valuate the contribution an integrated nitrogen

management strategy 12 could make to environmental protection;

12 An integrated nitrogen management strategy takes a holistic approach for managing N r

. I n the context of the nitrogen cascade, all N r anthropogenic

creation and destruction mechanisms and all N r uses are recognized. T he strategy should take account of syner gies 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 N r , while limiting overall adverse effects.8 3. I dentify additional risk management options for EP A’ s consideration; and

4. Make recommendations to EP A concerning improvements in nitrogen research to support risk reduction.

I n this report the C om mittee 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 EP A requires statutory authority to take regulatory action. H ow ever, 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
T

o address the four objectives of this study, the C ommittee completed the following activities:

1. T he C ommittee used the nitrogen cascade framework

to determine the major sources of newly created N r in the U.S. T he flows of N r 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 N r is lost to the environment. T he same process was employed for ener gy production but, since all of the N r formed during ener gy production is lost to the environment, the C ommittee identified the important ener gy producing sectors that contribute to N r formation.

2. T he C ommittee examined the fate of N r lost to the

environment, estimated the amount stored in dif ferent systems (e.g., forest soils), and tracked N r 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 C ommittee identified

the impacts N r has on people and ecosystem functions as it moves through dif ferent systems.

4. T he C ommittee identified actions that could be

taken based on available science and management practices to improve the integrative management of N

. T he C ommittee suggested ways in which each of these actions could be accomplished and estimated that together they could decrease N r losses to the environment by about 25%.

5. T he C ommittee identified research needed to improve

the scientific foundation to support specific N r risk reduction activities.

F our public meetings were held during the course of the study and briefings were presented to the C ommittee by: EP A’ s O ffice of Air and Radiation, O ffice of I nternational Affairs, and O ffice of Water; the U.S. D epartment of Agriculture’s Agricultural Research Service, C ooperative State Research, E xtension and E ducation Service, and the E conomic Research Service; and external or ganizations such as the E ner gy Research C entre of the N etherlands, E nvironmental D efense F und, I nternational P lant N utrition I nstitute, I owa State University , Live F uels, and the Soil and Water C onservation Society

. Additionally, the C ommittee invited scientists and managers from EP A, other federal agencies, states and localities, academia, non-governmental organizations, and the private sector to participate in an O ctober 20-22, 2008 workshop and meeting on N itrogen Risk Management I ntegration. T he purpose of the workshop was to receive public input on several subjects: the C ommittee’ s preliminary assessment of N r problems, consequences, and remedies, with emphasis on risk reduction; the C ommittee’ s quantitative estimates of attainable N r reductions; and mechanisms whereby the N r strategy might be enacted. T he C ommittee took this public input into consideration as it developed this report. Structure of the report T his report contains six chapters. T he report was developed for a multifaceted audience of scientists and policy makers and therefore the level of detail varies in dif ferent sections of the document. T he introductory chapter provides an overview of problems caused by excess reactive nitrogen and describes the study objectives and approach. C hapters 2-6 discuss how the C ommittee has addressed the four study objectives and present specific findings and recommendations. T he findings and recommendations corresponding to each of the study objectives are consolidated in C hapter 6.

n Study objective 1 (identification and analysis of the

problems nitrogen presents in the environment and linkages among these problems) is addressed in C hapters 2 and 3. C hapter 2 focuses on the sources, transfer , and transformation of reactive nitrogen in environmental systems. C hapter 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 C hapters 4 and 5. C hapter 4 reviews the implications for risk reduction strategies for reactive nitrogen. C hapter 5 discusses integrated risk reduction strategies.

n Study objective 3 (identification of additional risk

management options for EP A’ s consideration) is addressed in C hapters 5 and 6. I n C hapter 6, the C ommittee identifies specific management goals for reducing the loss of reactive nitrogen to the environment. T he C ommittee believes that these represent realistic near -term management goals that can be attained using current technology. H owever , the C ommittee emphasizes that further reduction beyond9 these goals will be needed to protect many N -sensitive ecosystems and to ensure that health-related standards are maintained.

T he C ommittee 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 C hapter 6 contains a section describing the need for a comprehensive program to monitor N r in the environment.

T hroughout this report there are boxes containing summary statements labeled “ F indings.” Attached to these findings are one or more specific “Recommendations” for actions that could be taken by EP A or other management authorities.1011 

# 2 Sources, Transfer, and Transformation of Nr in Environmental Systems

T he C ommittee was charged with identifying and analyzing, from a scientific perspective, the problems N r presents in the environment and the links among them. T his chapter addresses two aspects of the C ommittee’ s work. T he first aspect is the introduction of N r into U.S. environmental systems from fossil fuel combustion and from food production, and the second aspect is the fate of N r after it is emitted to the atmosphere by fossil fuel combustion or lost to the air , water and soils from agricultural production systems. T he N r budgets and calculations for the U.S. exclude Alaska and H awaii because data were not available for these areas. Most of the N r introduced into the environment comes from the H aber -Boesch process. H aber -Boesch contributes three to four times the N r introduced from fossil fuel combustion and most of the H aber -Boesch introduction of N r comes through agriculture.



### 2.1. Nr Flux in the Environment As stated previously, although N i s a major required nutrient that governs growth and reproduction of living organisms, N r losses to the environment from human sources have a profound effect on air, water and soil quality. H um an consumption of energy to sustain economic development results in emissions of NO x to the atmosphere via fossil fuel combustion. C on sumption of food to meet nutritional requirements of a growing population results in agricultural emissions of NH 3 , urban and industrial emissions of NO x and N 2 O , and losses

o f NO 3 - and other N c ompounds to water bodies due to leaching and runoff. O nc e released into the atmosphere by either human or natural processes, these N r 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 N r can also undergo transformation mediated by a wide range of biotic and abiotic processes. T he introduction of N r into agroecosystems provides much of the world’s food. T he losses of N r to the environment throughout the food production process and during fossil fuel combustion contribute to many of the major environmental problems of today. T he impacts of N r on humans and ecosystems are discussed in the E xe cutive Summary and C ha pter 3 of this report. Some key issues concerning management of N r in the U.S. environment are the introduction of new N r by imports, fertilizer production, cultivation-induced biological nitrogen fixation ( C -B NF ), and fossil fuel combustion. O th er important issues are the distribution of N r within agricultural systems and populated systems and redistribution of N r through losses from those systems to the environment ( F ig ure 1). N at ional-level values for N r flu xes are displayed in T ab le 1. F lu xes that represent the introduction of new N r into the U.S. are marked with an asterisk and illustrated in F ig ure 1. I n specific sections of this report these values have been used to more clearly determine the flux and fate of N r in the U.S. F ig ure 2 I ll ustrates sources of N r 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 Industry12 *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* %

N 2 O-N emissions 1 0.8 8 Agriculture - livestock (manure) N 2 O-N 0.03
Agriculture – soil management N 2 O-N 0.5 Agriculture - field burning agricultural residues 0.001 Fossil fuel combustion - transportation* 0.1

Miscellaneous 0.1

NH x -N emissions 2 3.1 31
Agriculture: livestock NH 3 -N 1.6 Agriculture: fertilizer NH 3 -N 0.9
Agriculture: other NH 3 -N 0.1 Fossil fuel combustion – transportation * 0.2 Fossil fuel combustion - utility & industry * 0.03

Other combustion 0.2

Miscellaneous 0.1

NO x -N emissions 2 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 deposition b 6.9 16 Organic N 3 2.1 Inorganic NO y -N 4 2.7
Inorganic-NH x -N 4 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 production 9 6.0 14

Human waste N 10 1.3 3

Total Terrestrial inputs 43.5 100 Nr inputs to the **Aquatic** environmental system

Surface water N flux 11 4.813 Table 1 Notes

a. T he N r estimates in this table are shown with two significant digits or 0.1 million metric tons N per year (or T g 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. O btaining quantitative estimates of each of the N r terms and the associated uncertainties remains a major scientific challenge.

b. Reducing the uncertainty in total deposition of atmospheric N r 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) T g N /yr total anthropogenic N r deposition to the entire 48 States. T he EP A sponsored C ommunity Multiscale Air Quality ( C MAQ) Model run yielded a value of 4.8 T g N /
yr

. T he value shown for the total (6.9 T g N /yr) reflects the assumption that or gano-nitrogen species should be added to the model estimate as 30% of the total.

c. T he synthetic N total includes 5.8 T g N of fertilizer net imports to the United States (8.25 T g N imported - 2.41 T g N exported) plus 9.4 T g N of fertilizer produced in the United States in 2002.

T

able 1 Data Sources (all data reflect N fluxes in the United States in 2002)

1 E missions, N 2 O

- N (U.S. EP A I nventory of U.S. Greenhouse Gas E missions and Sinks, 1990-2006)

2 E missions, NH x - N ; E missions, NO x - N (U.S. EP A N ational E missions I nventory , release version O ctober 2007)

3 Atmospheric deposition, organic N (30% of total atmospheric N deposition, N ef f et al. 2002)

4 Atmospheric deposition, inorganic NO y - N and NH x - N (U.S. EP A C MAQ model)

5 N 2 fixation in cultivated croplands (US D A census of agriculture 2002, literature coefficients)

6 N 2 fixation in non-cultivated vegetation (unpublished data estimate after C leveland and Asner, 1990)

7 N et N imports in commodities and fertilizer trade ( F

A O - F A O S T A

T )

8 Synthetic N fertilizer use ( F A O - F A O S T A

T and Association of American P lant F ood C ontrol O fficials - AA PFCO )

9 Manure N production (US D A census of agriculture, literature coefficients)

10 H uman waste N (U.S. C ensus Bureau population census, literature coefficients)

11 Surface water N flux (USGS S P ARR O W model, after Alexander et al., 2008) 

### 2.2. Sources of New Nr to the Environment C reation of “new” N r in the environment refers to N r that is either newly fixed within or transported into the United States. T his “new” N r highlights where N r is introduced into ecosystems. N ew N r arises from fossil fuel combustion, food production, and materials production ( T able 1).

F ossil fuel combustion emits N r (mostly NO x ) to the atmosphere. F ossil fuel combustion introduces 3.5 T g N /yr and 1.9 T g N /yr of NO x - N to the atmosphere from transportation, and utility/other industry sources, respectively ( T able 1). Another 0.2 T g N /yr of NH 3 - N and 0.1 T g N /yr of N 2 O

- N is emitted from the same sources ( T

able 1). T hus the total amount of N r created by fossil fuel combustion is 5.7 T g N /yr , of which > 90% is in the form of NO x - N

. Sy nthetic N r fertilizers are typically produced by the H ab er-Bosch process and used primarily in agriculture to support food production. P ro duction of fertilizers within the U.S. introduces N r into U.S. terrestrial landscapes at the rate of 9.4 T g N /y r, and net imports of fertilizer via world trade introduce 5.8 T g N /y r. O f th is total (15.2 T g N /y r), 9.8 T g N /y r is used as fertilizer on farms and 1.1 T g N /y r 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 T g N /y r is introduced for non-fertilizer uses, such as for production of plastics, fibers, resins, and for additives to animal feed ( T ab le 1).
Additional N r is introduced into the U.S. from C -B NF by agricultural legume crops such as soybean and alfalfa (7.7 T g N /yr), and from imports of N contained in grain and meat (0.15 T g N /yr) ( T able 1).

T hus in 2002, anthropogenic activities introduced a total of 29 T g N into the U.S., mostly in support of food production, although turf production, industrial uses and fossil fuel combustion were also important sources. N atural sources of N r in the U.S. are biological nitrogen fixation (B NF ) in unmanaged landscapes, and lightning. T he former contributes 6.4 T g N /yr ( T able 1) and the latter 0.1 T g N /yr

. C learly , anthropogenic activities dominate the introduction of N r into the U.S.
Losses of N r to the environment in the U.S. occur during fossil fuel combustion and food production. T he former occurs immediately , as N r formation during combustion is inadvertent and the N r , primarily as NO x , is emitted directly into the atmosphere. T he latter occurs through all stages of food production and consumption. T he following four Sections (2.2.1-2.2.4) of this report document the magnitude of the losses of N r to the environment from the various components of both ener gy and food production.14 2.2.1. Nr Formation and Losses to the Environment from Fossil Fuel Combustion F ossil fuels such as coal, petroleum, and natural gas provide about 80% of all ener gy production in the U.S. (based on year 2000). When these fuels are burned at high temperatures, NO x is formed. T he source of N is either the N contained in the fossil fuel or the N 2 that makes up about 80% of atmosphere. F uel-derived N is important in the case of burning coal (which contains N ), while atmospheric-derived N 2 is transformed to NO x during higher temperature processes that occur when gasoline or diesel fuel is burned in motor vehicles ( T able 1). As F igure 3 indicates, in the U.S., highway motor vehicles account for the lar gest anthropogenic source of NO x (36%), followed by off-highway vehicles, electric utilities, and industrial processes. E missions from aircraft make up only about 1% of the U.S. total for NO x , but a large proportion of this is released in free tropospherewhere lifetimes are long and adverse impacts wideranging. As such, continued reductions are encouraged (e.g., EP A Regulatory Announcement: N ew E mission Standards for N ew C ommercial Aircraft E ngines, www

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

F ig ure 3 also illustrates that the amount of NO x (reported as metric tons of N ) released from various fossil fuel sources has decreased dramatically from 1970. T ot al 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. O ve rall this represents a decrease of over 30%. T he top sources (highway vehicles, off-highway vehicles, electric utilities, and other industrial and combustion systems) show decreases of 15-30% from 1990 to 2002 ( F ig ure 4). Reductions were the highest for “other” systems followed by electric utilities. T he se decreases are most likely the result of changes in regulations and control technologies for these stationary systems. More recent preliminary information provided by EP A indicates that electric power NO x emissions may have decreased 70% between 1990 and 2009 and that the electric power sector’s NO x emissions now account for about 12% of anthropogenic NO x emissions in the U.S. (U.S. EP A, 2010b). T o 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%. T hi s decrease however, is accompanied by an increase in miles traveled, which suggests that the actual decrease in a single vehicle is larger. O ff -highway vehicles showed an increase in emissions, potentially due to better quantification of these sources. Such sources include locomotives and marine engines. EP A is in the process of implementing a number of regulations that will reduce NO x emission from mobile sources (see Appendix F ). F ig ure 5 (provided by EP A) projects decreases in U.S. mobilesource NO x emissions. T he implications of these recent regulations are not reflected in the quantitative analyses presented in this report. H ow ever, additional control of these and other sources could further decrease emissions. I n fact, technological development in the locomotive industry shows that decreases of approximately 70% are possible. F ur ther decreases would require more innovative, expensive methods such as selective catalytic reduction

Data are reported as thousands of metric tons of N converted from NO x as NO 2 Data source: www.epa.gov/ttn/chief/trends. More recent information provided by EPA indicates that electric power NO x emissions decreased 70% between 1990 and 2009 (U.S. EPA, 2010b). Figure 3: U.S. NO x

emission trends, 1970-2006Tg of Nitrogen (converted from short tons of Nox, as NO2)

Figure 4: Percent reductions in NO x

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 NO x emissions may have decreased 70% between 1990 and 2009.15 (S C R) with urea injection. E ng ine manufacturers are also investigating using S C R systems for diesels. H ow ever, it must be noted that these systems emit small amounts of NH 3 and must be operated properly to avoid trading off NO x emissions for NH 3 emissions.

I t should also be noted that it is difficult to control nitrogen emissions by regulating one source solely. As seen in T ab le 2 (data taken from 2001 for illustrative purposes have been reported with a higher degree of precision than is likely to be known), in T ex as, fuel combustion sources are on the same order as emissions from highway vehicles. By comparison, in C al ifornia, highway and off-highway vehicles are the dominant source of nitrogen emissions (over 75%). T he se results can be attributed to differences in the two states’ industries and power plants. I n T ex as,almost 40% of the power generation is from coal-fired plants. O n the other hand, C al ifornia 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%). T ab le 2 also shows emissions in F lo rida, O hi o, and I ll inois. T he emission of NO x from highway vehicles is likely related to population. F or example, the estimated population of C al ifornia in 2006 was 36.4 million people versus O hi o and I ll inois 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 N r and accounts for more N r losses to the environment than any other economic sector. Synthetic fertilizers are the largest sources of N r input to agricultural systems. T he next largest source is N fi xation in cultivated croplands ( T ab le 1). T he major pathways by which N r is lost from these systems include NO 3 - losses from leaching, runoff and erosion and gaseous emissions via volatilization of NH 3 and NO x and nitrification/denitrification. Similar loss pathways occur for N r that cycles through livestock systems, which also account for a large portion of N r flu x (predominantly as NH 3 ) in animal agricultural systems (Aneja et al., 2006). T he refore, assessment of N r 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 NO x

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 finalrule, 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 NO x

emissions based on 2001 data, and tons of NO x

as NO 2



### 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**

**2014 2020**  **2030 2009**  **2002**16 of N r fluxes in both crop and livestock systems, and
t he trends in management practices that have greatest influence on N r 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 ef fect on agriculture both nationally and globally by increasing crop yields. H owever , the high loading of N r from agricultural nutrient sources has led to deleterious ef fects 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 O bt aining accurate data on fertilizer use is a critical first step in understanding N r cycles in agriculture. T he re 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 F er tilizer T on nage Reporting System (U FT RS ) 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. T he Association of American P la nt F oo d C on trol O ffi cials (AA PFCO ) tallies and publishes the statewide fertilizer sales data annually ( T er ry and Kirby, 2006). I nf ormation published by AA PFCO i s a widely used source of fertilizer use data. I t is typically assumed that fertilizers are used in the same region in which they were sold. T he annual data published by AA PFCO
, wh ich are based on commercial fertilizer sold (and taxed in some but not all states), is the only state-level data source available. T hi s state-level data source includes fertilizer sales for both agricultural and non-agricultural
purposes. T he se 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 (p ersonal communication, Stan D ab erkow, US D A- E RS ).

T he U.S. D ep artment of Agriculture (US D A) N at ional Agricultural Statistics Service ( N AS S) 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. N AS S periodically collects fertilizer, pesticide, and pest management data from a stratified random sample of farmers at the field level. T he N AS S report represents a useful data source but also would require extrapolation across reported crop acreage to represent a complete sample of application rates. T he U FT RS was not designed to track the source of inor ganic nutrients applied to agricultural land on the geographic scale needed for watershed modeling. T he system only tracks sales of synthetic fertilizers and not manure or biosolids applied to farmland. I n addition, geographical data associated with each sale may or may not be near the actual point of application. H owever , 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. T hose changes would help tar get interventions and extension programs to improve nutrient management and reduce nutrient losses. T he 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. I mprovements in the database format and web-based

access
3. T he identification of funding sources to support

development of a more accurate, accessible, and comprehensive database system

4. E ducation and outreach to improve precision of

reported fertilizer tonnage, including a clear distinctionbetween nutrients used in crop, livestock, and nonagricultural operations

I n 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.

T he C he sapeake Bay watershed provides a good example of the fertilizer data dilemma. While the fertilizer tonnage that is currently being utilized to calibrate the C he sapeake Bay P ro gram Watershed Model is relatively accurate, the county-specific tonnage may have an accuracy of only ±20 to 50% ( C he sapeake Bay Scientific T ec hnical Advisory C om mittee, 2007). F or 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 ( C he sapeake Bay Scientific T ec hnical Advisory C om mittee, 2007). I t is also possible that fertilizer reported for crop agriculture may actually have been used for lawn and turf, forestry, or other non-agriculture applications.

N itrogen fertilizer application data collected on a specific crop-by-crop basis that can be associated with17 crop yields and location are essential for assessing both use patterns and efficiency. As previously mentioned, the US D A- N ASS maintains a database on N fertilizer rates applied to the major crops (corn, wheat, cotton, soybeans, and occasionally other crops). T he data are gathered from farmer surveys conducted every other year

. T hese data represent a source of useful information ( Pr otocols for Farming Reporting , Mark R. Miller, US D A, N ASS).

D ata derived from N ASS farmer surveys include six categories: field crops, fruits and vegetables, nurseries/ floriculture, livestock use, and post-harvest application. F or each group, N ASS collects fertilizer , pesticide, and pest management data on a stratified random sample of farmers at the field level. O ne field represents an entire farm for each sample in the field crops survey

. F ruit and vegetable information are collected for the entire farm. I f the field chosen for sampling has had manure applied in conjunction with inor ganic fertilizer, only the inorganic portion will be reported because the survey does not ask about manure.

C ertain core crops are surveyed every other year on an even/odd basis and states are selected to cover at least 80% of planted acres. T here has been a recent five-year gap in the collection of N ASS data for corn and other commodity crops. N ASS was scheduled to resume coverage of corn and other commodity crops in 2010. T his is a critical data gap and it is a problem given the lar ge changes in corn price and production area during this period. T hose data are needed to make progress in assessing fertilizer use and efficiency for major crops in the U.S. US D A N ASS must resume their yearly data collection for commodity crops ( C hemical Use Survey). P otential 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 N ASS survey data, the US D A E conomic Research Service released a report on fertilizer use that provided data on fertilizer consumption and type of fertilizer used from 1960-2006 ( F igure 6) and types of fertilizers used ( T able 3) (US D A, 2008). T he 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: C rop agriculture receives 60% of U.S. annual new N r inputs from anthropogenic sources (9.8 T g from N fertilizer , 7.7 T g from crop B NF versus 29 T g total) and accounts for 58% (7.6 T g) of total U.S. N r losses from terrestrial systems to air and aquatic ecosystems. H owever , 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 N it rogen fertilizer use efficiency ( NF U E ) is critical because higher use efficiency leaves less N r emaining to create potential environmental problems. H er e and throughout this report we define NF U E a s the grain yield per unit of applied N
, which is the product of two parameters: (1) the proportion of applied N f ertilizer that is taken up by the crop, or N f ertilizer recovery efficiency [recovery efficiency (R E ) in kg N u ptake per kg N a pplied], and (2) the physiological efficiency with which the N
t aken 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 u ptake] ( C as sman et al., 2002) 13 . All else equal, when higher NF U E i s achieved without yield reduction, the crop takes up more of the applied N a nd incorporates it into its biomass, which leaves less of the applied N r at risk for losses via leaching, volatilization, or denitrification. F ix en (2005) reports that there is substantial opportunity for increasing NF U E t hrough 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.

I n most cropping systems, R E is the most important determinant of NF U E

. A recent review of R E for cereals based on field studies around the world, mostly conducted on “small-plot” experiments at research stations, reported mean single year R E values for maize, wheat and rice of 65%, 57%, and 46%, respectively (Ladha et al., 2005). H owever , crop R E values based on actual measurements in production-scale fields are seldom greater than 50% and often less than 33%. F or example, a review of R E in dif ferent cropping systems estimated average recoveries of 37% for maize in the north central U.S. ( C assman et al., 2002). I t 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

. F or maize in the U.S. C orn 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). T herefore, 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.
H owever , 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 R E is fundamental for prioritization of research and education investments, both in the public and private sectors. While management can substantially improve R E 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 ( F igure 6), yields of all major crops have continued to increase. Because crop yields are closely related to N uptake ( C assman et al., 2002), these trends imply a steady increase in NF U E and reduced N losses to the environment because more of the applied N is held in crop biomass and harvested grain. Greater NF U E has resulted from two factors. T he first factor is a steady improvement in the stress tolerance of corn hybrids ( D uvick and C assman, 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. T he 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

. E xamples 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). F or maize, which receives the lar gest share of total N fertilizer in the U.S. (44% in 2005), NF U E decreased markedly in the 1960s because N fertilizer rates rose more quickly than maize yields. H owever , 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 NF U E from 1974-76 to 2002-05 ( F igure 7). Similar T

able 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 NH 3 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. I f the yield response is more than half a bushel (12.7 kg) of corn per pound (0.45 kg) of N at the mar gin 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 i n 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

a pplication than might otherwise be appropriate for the N

u tilization in the four years of lower yield. T hi s presents a real dilemma if the policy goal is to reduce N t ransfers to the environment, especially in the four years of average or lower yields. Meeting this challenge will requireapproaches such as the development of real-time, inseason, decision-making tools that allow crop producers
to use N f ertilizer rates for average yields at planting and during early vegetative growth, and a final top-dressing as required to meet any additional N d emand above this amount due to favorable climate and soil conditions that support higher than average yields ( C as sman, 1999; C as sman 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 f ertilizer. 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 i nputs from legume N fi xation and recycling of N i n manure and compost. At issue, however, is whether such systems actually reduce N r losses to the environment because the same loss mechanisms and pathways operate

on N f rom 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 f ertilizer. Lower yields would require more land in production to meet food demand. T he refore, a key issue is whether the tradeoff in reduced N f ertilizer inputs to more diverse crop rotations with organic N i nputs would actually result in less N r 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 N r losses from agriculture would be to shift and/or adjust cropping systems across the landscape. T his would involve changes in land use as well as crops. T his approach for parametric reductions in nitrogen was analyzed extensively in the Gulf of Mexico H ypoxia Assessment ( D oering et al., 1999). As part of a modeling exercise to address opportunities and consequences for reducing N r , crop rotations as well as tillage practices and fertilizer inputs were adjusted to meet successive constraints on excessive N r while also maximizing consumer and producer welfare to the extent possible. T he model favored those crops and cropping practices that had lower N r 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 N r reduction scenario, crop acreage was reduced by about 6%. T his 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 C rop 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

. O ne way to decrease leakage is to apply less N fertilizer to croplands. F or example, H u et al. (2007), using the SW A

T model, predict that decreasing N fertilizer application rates by 10 to 50% of those used in the 1990s in the upper E mbrarras River watershed in east central I llinois would decrease NO 3 - output to the river by 10 to 43%. T his 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.

C ropping 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 ( C assman et al., 2002), there would be a proportional decrease in crop yields in response to a decrease in the amount of N fertilizer application. T he 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. H ence, 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 of fset the reduction in the amount of applied N

. I t 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 or ganic matter over time. T hus 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 f ertilizer input to cereal crop production is the effect on crop quality. I n D en mark, for example, national efforts to decrease N r losses to the environment resulted in a large reduction in use of N f ertilizer in crop production ( F ig ure 8). T hi s situation occurred as part of D en mark’s21 response to the E ur opean Union N it rate D ir ective. Synthetic fertilizer N u se in D en mark decreased from approximately 0.4 T g N i n 1991 to 0.2 T g N i n 2002. Animal manure N a pplication decreased from 0.25 T g N t o approximately 0.24 T g N d uring this time period. D es pite this large decrease in N f ertilizer 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 ( F ig ure 9). Grain protein content in wheat is critical for determining its quality for bread. I n the U.S., a grain protein content of 12% is considered the threshold for good quality bread wheat, and N f ertilizer application rate has a large influence on determining this trait ( C as sman et al., 1992). As can be seen, grain protein content has declined from 12 to 10% in D en mark 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)? D o these costs offset the environmental benefits created by decreasing N fl ows 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 f ertilizer input if the methods used to apply N a re modified to improve NF U E

a s illustrated by U.S. maize yield trends and fertilizer use ( F ig ure 7). H ow ever, at some point continued yield increases will require additional N i nput because of the tight relationship between yield and crop N u ptake requirements. T he goal of N f ertilizer management should therefore be to achieve highest possible NF U E

w hile also sustaining crop yield increases adequate to meet food demand without need for a large expansion of crop production area. I n fact, studies that evaluate NF U E i n production agriculture find a wide range in NF U E

a mong farmers. F or farmers who achieve high NF U E
, a reduction in N f ertilizer rate would reduce yields. H en ce, mandating an across-the-board reduction in N f ertilizer use penalizes the good farmers who achieve high NF U E

a s their yields would likely fall while it would have little effect on yields of farmers who currently over fertilize ( C as sman et al., 2003; D ob ermann and C as sman, 2004).

F in ally, if yields were significantly reduced as a result of lower N f ertilization 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 C on servation Reserve P ro gram ( C R P ). Such conversion would have two negative results. F ir st, additional N l osses from these acres would occur due to relatively low N f ertilizer efficiency that typically occurs on marginal land that has multiple soil constraints to crop growth and yield. Second, conversion from C R P t o 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: N r 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 f ertilizer uptake efficiency. T hi s 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 ( C as sman, 1999). More diverse cropping systems with decreased N r fertilizer input may also provide an option to increase NF U E o n a large scale if the decrease in N r 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. H ow ever, crop cultivars and agronomic practices are changing rapidly, which changes N r equirements, but current efforts in research, extension, and conservation
programs on N m anagement within these rapidly evolving systems are not adequate to meet the challenge of providing better information to increase NF U E

. Recommendation 2: To obtain better information on Nr inputs and crop productivity, the Committee recommends that:

Recommendation 2a: Data on NFUE and N massbalance, 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 B NF i n 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 b ecause N r from B NF i s prone to the same loss pathways as N r from commercial fertilizers. T o quantify B NF d ue to human cultivation of crops, the C om mittee calculated the annual agricultural fixation for 2002 using crop areas and yields reported by the US D A C en sus of Agriculture (US D A, 2002). T he C om mittee multiplied the area planted in leguminous crop species by the rate of N fi xation specific to each crop type, assigning rates based on a literature review, as summarized in T ab le 4 below and shown relative to other inputs in T ab le 1. Annual nitrogen inputs to cropping system from B NF b y legume crops was 7.7 T g N /y r in 2002, accounting for ~15% of the overall N r in puts to the terrestrial landscape from all sources and 20% of the agricultural sources ( T ab le 1). Soybean and alfalfa contributions are the most important agricultural legumes in terms of nitrogen input and contribute 69% of total B NF i nputs in U.S. agriculture.

Emissions factors and losses to the environment from fertilizers and organic nitrogen sources Agriculture is a significant contributor of N r inputs into the atmosphere. N itrogen fertilizer losses vary greatly

T

able 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 ( C assman et al., 2002). I n addition, any factor that af fects crop growth vigor and root system function also affects the ability of the plant to recover applied N efficiently

. F or example, denitrification can range from 0% to 70% of applied N (Aulakh et al., 1992). T his 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.

D es pite this variation, watershed, regional, and national assessments of carbon and N c ycling often rely on average values for losses from each pathway. F or example, the I nt ergovernmental P an el on C li mate C ha nge ( IPCC ) assumes that 1% of applied N f ertilizer (uncertainty range of 0.3-3.0%) is lost from direct emissions of N 2 O at the field level due to nitrification/ d enitrification. T hi s assumption is based on analysis of all appropriate scientific publications that report these losses for specific crops and cropping systems ( IPCC
, 2007a). T he same 1% default emission factor for field-level N 2 O emission is applied to other N i nputs from crop residues, organic amendments such as manure, and from mineralization of native soil organic
matter. D at a from scores of field studies were used to obtain this average value. A number of recent studies confirm that N 2 O losses to the environment during the
g rowing season at the field level can represent <1% of the applied nitrogen – even in intensive, high-yield cropping systems (Adviento-Borbe et al., 2006). D es pite these average values, it is also clear that N 2 O losses

c an vary widely even within the same field and from year to year due to normal variation in climate and crop management ( P ar kin 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). P re dicted increases and decreases in rainfall will likely have a dramatic impact on nitrogen export from agricultural fields. F or example, precipitation is predicted to increase in the upper Mississippi watershed and, other factors being equal, N

e xport should increase (e.g., Justic et al., 1995a,b).
Additional indirect N 2 O emissions result from d enitrification of volatilized NH 3 deposited elsewhere or from NO 3 - lost to leaching and runoff as the N r

ca scades through other ecosystems after leaving the field to which it was applied. H er e the IPCC a ssessment protocol assumes that volatilization losses represent 10% of applied N
, and that N 2 O emissions for these losses

a re 1% of this amount; leaching losses are assumed to be 30% of applied nitrogen and N 2 O emissions are 0 .75% of that amount ( IPCC
, 2007a). T he refore, the IPCC d efault value for total direct and indirect N 2 O

em issions represents about 1.4% of the applied N f rom fertilizer. By the same calculations, 1.4% of the N i n applied organic matter, either as manure or compost or in recycled crop residues, is also assumed to be emitted as N 2 O . Recent work funded by EP A used the D A YCENT

mo del to estimate N 2 O emissions from cropping systems ( D el Grosso et al., 2005). H ow ever, due to the cost of field validation of such models, there are relatively few validations across a representative range of cropping systems and environment. T he refore, it is not clear that use of such a complex model gives better estimates of N 2 O emissions than the more straightforward IPCC

a ssessment protocol.

O thers have estimated higher average global N 2 O losses of 3-5% of applied nitrogen fertilizer based on historical changes in atmospheric N 2 O content and changes in N r production during the past 50 years ( C rutzen et al., 2008), as opposed to the field-based estimates that form the basis of IPCC estimates. Because N 2 O is such a potent G H G, and given the more than two-fold difference in estimates of N 2 O losses, there is a critical need to improve understanding and prediction of N 2 O losses from agricultural systems. N 2 O emissions in the U.S. are estimated to be 0.78 T g N /yr ( T able 5) (U.S. EP A, 2005b).

Bi ogenic NO x 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; V el dkamp and Keller, 1997; C iv erolo and D ic kerson, 1998). H ow ever, NO x emissions by agricultural burning are relatively unimportant. Ammonia volatilization of N f rom applied fertilizer can be the dominant pathway of N l oss in rice soils and can account for 0% to more than 50% of the applied N d epending on water management, soil properties, and method of application (citations within P eo ples et al., Table 5: N 2 O 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 ( P eo ples et al., 1995, 2004). T he IPCC ( 2007a) uses a value of 10% of synthetic fertilizer N a pplication and 20% of manure N a s estimates of average NH 3 volatilization.

T

ak en together, N l osses from all forms of direct gaseous emissions from crop production systems can represent a substantial portion of applied N f ertilizer when soil conditions favor such emissions and there is a lack of synchrony between the amount of N a pplied and the immediate crop demand (Goulding, 2004). T he refore, achieving greater congruence between crop demand and the N s upply from fertilizer is a key management tactic to reduce N l osses from all sources. Success in reducing N l osses and emissions from agriculture will depend on increased efforts in research and extension to close gaps in our understanding of N c ycling and management in crop production, especially as systems further intensify to meet rapidly expanding demand for food, feed, fiber, and biofuel. Finding 3: N itrous oxide emissions from the N r inputs to cropland from fertilizer , manure, and legume fixation represent a large proportion of agriculture’s contribution to G H G emissions, and the importance of this source of anthropogenic G H G will likely increase unless NF U E is markedly improved in crop production systems. D espite 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 T he enormous use of liquid fuels in the U.S., the rising demand for petroleum based liquid fuels from countries like C hi na and I nd ia, and the decline in petroleum discovery all contributed to the recent record high petroleum prices. I n addition, most of the world’s petroleum reserves are located in politically unstable

areas. T hi s has provided strong motivation for policies promoting investment in biofuels made from corn, oil crops, and ultimately from cellulosic materials. I n th e 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). T he renewable fuels standard in the 2007 E ne rgy I nd ependence and Security Act ( EI SA ) 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. P ro duction of biodiesel from vegetable oils also is encouraged in EI SA , 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.

I n 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. H ow ever, EI SA 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 ( D oe ring and T

yn er, 2009). T he higher corn prices of 2007 and 2008 resulted in more land being planted to corn and higher N f ertilizer requirements. C or n 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. T hi s strong response to high demand for biofuel feedstock has led to concern about increased pressure on the environment from biofuels. O ne important factor is the increased N n ecessary for growing corn and cellulosic materials (Robertson, et. al. 2008). Biofuels production and consumption will also result in NO x emissions to the atmosphere. E xp ansion of corn or cellulosic materials production into marginal lands can be even more problematic with respect to nutrient leaching and soil erosion. C ha nges in N f ertilizer prices add uncertainty to the additional amounts of N t hat may ultimately be used in biofuel feedstock production. P ro duction 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 i n cattle manure (Klopfenstein et al., 2008).

I n F eb ruary 2010, EP A released its final rule for the expanded Renewable F ue ls Standard (R F S2 ) regarding greenhouse gas emissions from various types of biofuels based on life-cycle analysis as required by the 2007 EI SA . 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 commercial25 scale. I n contrast, life cycle analysis (L C A) 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 EI SA , 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

. D istillers grains are changing animal diets and af fecting N recycling in livestock. Both have important consequences for ef fective future management of N r , but current models and understanding of how expanded biofuel production will affect N r 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 I n the U.S., domestic animals produce 6.0 T g N /yr in manure and are the lar gest source of atmospheric NH 3 - N (1.6 T g N /yr) ( T able 1). Livestock also contribute to N 2 O

- N emissions, though in much smaller proportions (~4% of total U.S. N 2 O

- 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. T he production of chicken broilers increased more than four fold from 1970 to 2006 ( F ig ure 10) and milk production increased by nearly 60% in this time period ( F ig ure 11). T ur key production doubled and pork production increased about 25%, while meat from cattle (beef and dairy) remained constant ( F ig ure 10). Another trend in animal production has been for fewer animals to produce more animal products. F or example, the 60% increase in the amount of milk produced in 2006 compared to 1970 required 25% fewer cows ( F igures 1 1 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. T his 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 Reports26 greater growth rates of animals producing more meat in a shorter amount of time. I n 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 (US D A, N ASS, 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. T here were only 7% as many swine operations and 1 1% as many dairy operations in 2006 as there were in 1970 ( F igure 13). T here were half as many beef operations in 2006 as in 1970, but beef operations also expanded in size while smaller producers held jobs of f 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. T herefore, unless the manure is applied over a lar ger 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 T he trends in livestock production have both positive and negative environmental impacts. O ne 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. T his ef fect is partly the result of effectively reducing maintenance requirements during production. T he requirements for feeding animals can be divided into two components: maintenance and production. T he maintenance component is the feed that is used to keep the animal alive and healthy so that production is possible. T he production component includes the feed that is converted to animal protein and waste due to the inefficiencies of these conversions. T he maintenance component depends upon the number of animals, each animal’ s mass, and the time the animal is on feed. T hus, the maintenance requirement is diluted by faster growth rates and greater body weight at slaughter

. T he 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.

P ublic concerns about the potential environmental and health ef fect of air emissions from C A FO s expand the impacts of food production beyond those associated with traditional agricultural practices ( N R C , 2001; Aneja et al., 2009). I ncreased 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). T hese compounds are then transported by wind and returned to the surface by wet and dry deposition processes, which may have adverse ef fects 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 Reports27 Adverse effects, further discussed in C hapter 3 of this report, include eutrophication, soil acidification, loss of biodiversity, and reactions that increase the mass concentration of atmospheric aerosols ( P M 2.5 ). Aerosol formation occurs when HNO 3 reacts with basic compounds, and NH 3 reacts with compounds. E cosystem acidification can occur when HNO 3 is deposited from the atmosphere. I n addition, acidification can also occur when NH x is deposited due to the production of HNO 3 from nitrification via soil microbes. Soil acidification occurs when HNO 3 or NH 4 + deposits on soils with low buffering capacity and this can cause growth limitations to sensitive plant species. D eposition of NO 3 - or NH 4 + also causes eutrophication (i.e., an over-abundance of nutrients), which can promote harmful algal growth leading to the decline of aquatic species. I n fact, volatized NO 3 - can travel hundreds of miles from its source, affecting local and regional biodiversity far from its origin (Aneja et al. 2008b; James, 2008).

T he potential for reduced environmental impact from N r in livestock systems depends on the proportion of the total intake attributable to maintenance costs. T he commonly used tables for diet formulation published periodically by the N R C 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 ( N R C , 1994) but protein requirements were not divided between maintenance and production. F or example, a dairy cow producing 40 kg milk per year would divert about 25% of its ener gy and 12% of its protein to maintenance ( N R C , 2001).

I n 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. I n other words, the actual nutritional maintenance cost of a herd or flock is greater than it is for productive individuals only

. F or example, milk production requires non-lactating cows and heifers in the herd which do not produce milk but which consume nutrients. T hese additional maintenance costs are lower for broiler flocks than for cattle. Finding 5: T here are no nationwide monitoring networks in the United States to quantify agricultural emissions of greenhouse gases, NO
, N 2 O , reduced sulfur compounds, volatile organic compounds ( VOC s), and NH 3 . 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. I n contrast there is a lar ge network in place to assess the changes in the chemical climate of the United States associated with fossil fuel energy production, i.e., the N ational Atmospheric D eposition P rogram/ N ational T rends N etwork ( N A DP / NTN ), which has been monitoring the wet deposition of sulfate (S O 4 2- ), NO 3 - , and NH 4 + since 1978.

Recommendation 5: The status and trends of gases and particulate matter precursors emitted from agricultural emissions, e.g., NO 3 - and NH 4 + 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 F ro m 1970 to 2006, several feeding practices were changed for diets fed to livestock ( N R C
, 1977, 1988, 1996, 2001). I n 1989 and 1996, the N R C i ntroduced 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. F ee ding 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. F or poultry and swine, manufactured amino acids were added to diets, decreasing the need for protein by 30%. T od ay, 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% ( N R C
, 2001). P hy tase 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.

I t is difficult to estimate the combined ef fects of these changes in feeding practices, but we have calculated changes in manure N by using US D A N ASS data and assuming improvements in both production rates and ration formulation ( T able 6). I n the case of beef cattle diet formulation, the changes in feeding practices were determined by comparing the N R C 1976 recommendations with the N R C 1996 recommendations. Surprisingly
, N R C 1996 recommended greater total crude protein compared to N R C 1976 despite formulating for rumen degraded and un-degraded protein and considering amino acid content. T herefore, improved diet formulation did not decrease N intake for beef in this time range but the ef fect of reduced maintenance did improve efficiency of N utilization. Reduced nitrogen excretion from increased efficiency N itrogen excretion as fraction of animal production decreased from 1970 to 2006 ( T able 6). H owever , in cases where the total amount of animal production in the U.S. increased substantially (e.g., broilers), total N excretion increased. T he decrease in N excretion per unit of animal28 productivity was estimated by calculating the effects of changes in feeding practices and reduction of maintenance as described previously. T he data in T able 6 indicate that there has been an increase in N utilization efficiency for livestock products.

F or broilers, data are available to more accurately estimate the effect of changes in feeding and genetics

on N e xcretion over time. H ow ever, these data do not represent the time period of interest in this report. H av enstein 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 i ntake and production data, there was a 51% reduction in N e xcreted between these diets (Kohn, 2004). Similarly, Kohn (2004) compared N excreted by U.S. dairy cows in 1944 and 2001. I n 1944, the historically lar gest herd of dairy cattle in the U.S. (25 million cows) produced an average of 7 kg milk per cow per day (US D A N ASS, 2007). I n 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 T g N to 1.2 T g N , respectively

. T

able 7 provides information on manure production from animal husbandry in the U.S. F or T able 7, manure N was calculated for all U.S. animal agriculture using data on animal production from the 2002 C ensus of Agriculture (US D A, 2002). F or 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 C ensus of Agriculture, but using the updated information from the 2002 C ensus of Agriculture. F or 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 ( T

A N ), temperature, wind speed, p H
, 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).

EP A estimates annual manure N excreted in livestock production in the U.S. for the Inventory of U.S. Gr eenhouse Gas Emissions and Sinks (U.S. EP A, 2007e). F or the year 2002, these estimates indicate that a total of 6.8 T g of N was excreted in livestock manure. O nly a fraction of this N (~1.24 T g) was recovered and applied directly as a nutrient source for crop production. Approximately 1.8 T g N was transferred from the manure management systems, most likely by ammonia volatilization. O ther loss vectors include leaching and runof f during treatment, and storage and transport before soil application. T he remainder of the N was deposited in pastures and rangeland or in paddocks. T his N is also T

able 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 BM P s and engineered solutions it may be possible to reduce the emissions and dischar ge of odors, pathogens, and nitrogen compounds from agricultural operations (Aneja et al., 2008b,d).

T

ot al manure production reported in T ab le 7 in the contiguous U.S. was estimated using US D A’ s method and yields an estimate of 6.0 T g N /y r; while EP A’ s G H G inventory method in T ab le 8 yields a total for the U.S. of 6.8 T g N /y r in 2002. T he “greenhouse gas” method suggests 13% higher manure N p roduction. T hi s difference highlights uncertainty in the calculations. Although the values in T ab le 8 include Alaska and H aw aii, the values in T ab le 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: F ar m-level improvements in manure management can substantially reduce N r load and transfer. While the NPDE S permitting process for C A FO s does include nutrient and manure management plans to limit the transport of land applied nutrients off farms, substantial off-farm transport of N r still occurs, especially via atmospheric transport. T he re 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 T

urf grasses cover 12.6-16.2 million ha across the continental U.S. (Milesi et al., 2005). T he area under turf grass is roughly the size of the N ew E ngland states and occupies an area up to three times lar ger than that of irrigated corn ( T he Lawn I nstitute, 2007). T he majority of this turf area (approximately 75%) is in residential lawns. About 80% of all U.S. households have private lawns ( T empleton et al. 1998) that average 0.08 ha in size ( V

inlove and T orla, 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. N itrogen also may be derived from atmospheric deposition or recycled decomposition of soil and grass clipping or ganic 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. D epending 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 buf fers.

T

urf 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. P etrovic, personal communication, June 5, 2007). T he C ommittee has arrayed the dif ferent turf managements into the following three groups according to the estimated amount of N -fertilizer applied annually ( T able 9): residential lawns maintained by homeowners (0.73 kg/100 m 2 ), residential lawns cared for by professional lawn care companies (2.92 [range, 1.95-7.3] kg/100 m 2 ), and athletic fields and golf courses (3.89 [range, 2.64- 6.64] kg/100 m 2 ). T he estimate of total N -fertilizer used

T

able 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 T g N /year , or 9% of the total average annual N -fertilizer used between 1999 and 2005. D epending 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.

T

urf 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 NO 3 - can range between 0 and 50% of N applied ( P etrovic, 1990). N itrogen 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 p H

. I n 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).

N it rogen runoff losses are poorly quantified but a range similar to leaching is probable (A.M. P et rovic, personal communication). T he chemical form of fertilizer N d oes not impact leaching/runoff unless the fertilizer is applied in late autumn ( P et rovic, 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 N r to the atmosphere can be significant when urea is applied. Measured denitrification losses are usually small, but depend upon timing of N

a pplication relative to soil water status, irrigation, and temperature. T yp ically 25% of N a pplied is not accounted for in runoff, leaching, and uptake/removal, or soil sequestration (A.M. P et rovic, personal communication). T hi s suggests that volatilization and denitrification are important loss vectors. N it rogen 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 l osses than from well managed lawns ( P et rovic and Larsson-Kovach, 1996; P et rovic, 2004b), Guillard (2008) recommends not fertilizing lawns of acceptable appearance. F ur ther, 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 P etrovic (personal communication) half the lawns in the U.S. may not receive any fertilizer

. T hose lawns are presumably satisfactory to their owners. F urther 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

. T hese practices can potentially reduce N fertilization (and subsequent leaching risk) on turf by one third or more, saving 0.4 or more T g 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. E ven 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 m 2 /ha was used with application rates (kg N/100 m 2 ) of: 0.73 for nominal fertilization, 2.92 for professional lawn care, and 3.89 for high maintenance areas kg N/100 m 231 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
T his section discusses the transfers and flows of N r within and between environmental systems ( E S), which include atmospheric, terrestrial, and aquatic environments. T he first Section (2.3.1) contains information on N r deposition from the atmosphere to terrestrial and aquatic systems, presents estimates of input and recycling of N r within terrestrial systems, and discusses movement of N r from the terrestrial to the aquatic system. T he second Section (2.3.2) presents an estimate of storage of N r within the terrestrial system. Areas of uncertainty in N r transfer and transformation are discussed in Section 2.3.3. Section 2.3.3 also contains an example analysis of N r input and fate in 16 watersheds in the northeast U.S. N o comprehensive national data are available to assess the transfer and transformations of N r in and between the atmosphere, terrestrial systems and aquatic systems. T he example analysis in Section 2.3.3 shows how an evaluation of inputs and fate of N r could be conducted for a lar ge 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 N r content of terrestrial and aquatic ecosystems for the U.S., but the magnitude and mechanisms of N r 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 ( P aerl et al., 2002). O ther 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. V aligura 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%. NO x , NH 3 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 NO x , because of nonlinearities in the photochemistry, generate substantial amounts of tropospheric ozone (U.S. EP A, 2006a). T

otal 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 N r species. A portion of the N r deposited to the earth’ s surface is re-emitted as NH 3 , NO , or N 2 O (Galbally and Roy, 1978; Kim et al., 1994; C iverolo and D ickerson, 1998; IPCC , 2007a; C rutzen et al., 2008). Although naturally-produced N r is involved, anthropogenic N r dominates over most of the U.S. I n Appendix A we provide a review of the state of the science concerning the total annual N r deposition and trends in that deposition to the contiguous 48 states.

D eposition involves both oxidized and reduced N species. O f the oxidized forms of atmospheric N , all the members of the NO y family ( NO
, NO 2 , NO 3 , N 2 O 5 , HONO
, HNO 3 , NO 3 - , P A

N and other or gano-nitrates, R ONO 2 ) can be transferred from the troposphere to the surface, and some (e.g., NO ) under go bidirectional flux. V

olatile amines are also detected as NO y compounds (Kashihira et al., 1982; Wyers et al., 1993). Although a potent G H G, N 2 O is only emitted, not deposited and therefore has not been considered in the material presented here and in Appendix A. O f the reduced forms of atmospheric nitrogen, NH 3 and NH 4 + play a major role. T here is also evidence of deposition of or ganic 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. NO x budget, especially in summer (Keene et al., 2002; H orowitz et al., 2007; D uce 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. T he 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 N r in the U.S. and the relationship between emissions of N r and observed deposition.

Consideration of NO y

as a supplement for the current NO 2 National Ambient Air Quality Standard T he six principal (or criteria) pollutants for which EP A has established N ational Ambient Air Quality Standards ( N AAQS) include “oxides of nitrogen” (the sum of NO and NO 2 ) or NO x . T he specific chemical compound nitrogen dioxide ( NO 2 ) has been selected as32 the indicator for compliance with the N AAQS for NO x . T he levels of primary and secondary standards for NO 2 are identical at 0.053 ppm (approximately 100 μg/m 3 ) in annual arithmetic average, calculated from the one-hour NO 2 concentrations. I n 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 N r species. EP A’ s recent I ntegrated Science Assessment for O xides of N itrogen and Sulfur – E cological C riteria (U.S. EP A, 2008d) evaluated the scientific foundation for the review of the secondary (welfare-based) N AAQS for oxides of nitrogen and concluded that: The instrumentation deployed at pr esent in the routine monitoring networks for determination of gas-phase NO 2 and SO 2 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 NO 2 is inadequate to protect welfare, there is a straightforward technical fix to this problem. NO x ( NO + NO 2 ) is a variable, often small component of reactive oxidized nitrogen ( NO y ) as has been noted in EP A’ s I nt egrated Science Assessment for O xi des of N it rogen and Sulfur – E co logical C ri teria and extensively documented in the reviewed literature ( F ah ey et al., 1986; F eh senfeld et al., 1987; D od dridge et al., 1991; P ar rish et al., 1993; P ou lida et al., 1994; Ridley et al., 1994; E mm ons et al., 1997; Liang et al., 1998; Munger et al., 1998; Zhou et al., 2002; T ak egawa et al., 2003; P ar rish et al., 2004a,b; H or ii et al., 2004, 2006; D un lea et al., 2007; Kleinman et al., 2007; H ar grove and Zhang, 2008; Luria et al., 2008; Zhang et al., 2008; Schwab et al., 2009; Luke et al., 2010). T he standard chemiluminescence NO / NO x technique measures NO d irectly via reaction with O 3 , and NO x ( NO + NO 2 ) by conversion of NO 2 to NO

o n hot molybdenum. I t had been thought that NO x would make up nearly all of NO y in heavily polluted urban areas, but results including some from Mexico C it y ( D un lea et al., 2007) indicate that this is not the case. I nt erferences were up to 50% of the ambient NO 2 concentration even in a heavily urban area. Recent results from H ou ston, T X (L uke et al., 2010) showed a median NO x / NO y ratio of ~0.63 between 1300 and 1500 local time, a time of rapid ozone production. NO 2 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. H ow ever, given the inadequacy of NO 2 as a standard to protect welfare, the C om mittee recommends that EP A r eexamine the criteria pollutant oxides of nitrogen and the indicator species NO 2 and consider adding chemically reactive nitrogen as a criteria pollutant, and NH x and NO y as indicators to supplement the NO 2 N ational Ambient Air Q uality Standard.

T he references listed above also document that current suite of NO x monitors are useful for proving compliance with the current NO 2 standard, but suffer substantial interferences and the extent of these interferences varies with time and location. Reactive N c ompounds, especially HNO 3 , can be lost on the inlet components. Because the inlet does not transmit all reactive nitrogen compounds with high efficiency, commercial NO x monitors may provide an upper limit for NO x and a lower limit for NO y , but they measure neither NO x nor NO y precisely. N um erical simulations do not currently produce consistent results for the partitioning of NO y species, for example see Archibald et al., (2010). T he data from the current monitoring network are of limited value for determining exposure to NO 2 , for evaluating chemical transport models ( CT Ms ), or for assessing efficacy of emissions control strategies (Mc C le nny et al., 2002).

NO y monitors differ from the NO x monitors currently in use in the position of the hot molybdenum NO 2 to NO

c onverter and in calibration and operation (e.g., T he rmo Scientific Model 42i- Y ). I n an NO x monitor the converter is well downstream of the inlet, behind a filter that removes particulate matter, while in an NO y monitor the converter is at the inlet ( F eh senfeld et al., 1987; D od dridge et al., 1991; D od dridge et al., 1992; P ar rish et al., 1993; P ou lida et al., 1994; Ridley et al., 1994; E mm ons et al., 1997; Liang et al., 1998; Munger et al., 1998; P ar rish et al., 2004b; D un lea et al., 2007; H ar grove and Zhang, 2008; Luria et al., 2008; Schwab et al., 2009). T hu s the current monitors could be replaced or retrofitted to measure NO y with a detection limit of 0.1 ppb at reasonable expense for equipment and for training operators. NO y monitors, like NO x monitors, provide proof of compliance with NO 2 standards. F or specific measurement of NO 2 , commercial instruments for selective NO 2 reduction (e.g., Air Quality D es ign, I nc ., Wheat Ridge, CO ) or direct NO 2 measurement are beginning to be available ( P ar rish and F eh senfeld, 2000; C as tellanos et al., 2009) and these may offer an alternative approach. D ir ect monitoring of NO 2 and other N r species is a long-term goal, but new techniques will require thorough testing in polluted environments. NO y monitors using hot molybdenum compare favorably with other NO y techniques such as gold-catalyzed CO r eduction and several studies have concluded that NO y can be reliably measured in suburban and urban environments ( F eh senfeld et al., 1987; C ro sley, 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 N r to the earth’ s surface, but there are uncertainties in33 the characteristics and absolute magnitude of the flux. P ollutants not deposited are exported from the continent and alter the composition and radiative balance of the atmosphere on a lar ge scale. A review of the literature revealed the following major points concerning the present state of the science:

1. Measurements from the N ational Atmospheric

D eposition P rogram ( N A DP ) 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. T he reduced ( NH 4 + ) and oxidized ( NO 3 - ) 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. F or the U.S. east of the Mississippi River , dry

deposition data have also been analyzed – the C lean Air Standards and T rends N etwork ( C AS TNET ) monitors vapor phase HNO 3 , as well as particulate NO 3 - and NH 4 +

. T hese measurements indicate 7.75 kg N /ha/ yr total deposition (5.46 wet 2.29 dry) over the E ast. C onspicuous by its absence from this number is dry deposition of ammonia.

4. D ecreases in NO x emissions appear to have led to decreases in NO 3 - deposition. N A DP 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.

NO y budget indicates that about 70 % of the NO x 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 NO y include dry deposition of unmonitored members of the NO y 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 NO y 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 N r for the period 2000-2004 is ~1 1 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 N r remains high, but evidence is strong that atmospheric deposition of N r to the earth’s surface as well as emissions from the surface to the atmosphere contribute substantially to environmental and health problems. N it rogen dioxide, NO 2 , is often a small component of NO y , the total of oxidized nitrogen in the atmosphere. T he current N AA QS for NO 2 , as an indicator of the criteria pollutant “oxides of nitrogen,” is inadequate to protect health and welfare. NO y should be considered seriously as a supplement or replacement for the NO 2 standard and in monitoring. Atmospheric emissions and concentrations of N r from agricultural practices (primarily in the form of NH 3 ) have not been well monitored, but NH 4 + ion concentration and wet deposition (as determined by N A DP and NTN ) appear to be increasing, suggesting that NH 3 emissions are increasing. Both wet and dry deposition contribute substantially to NH x removal from the atmosphere, but only wet deposition is known with much scientific certainty. T hu s consideration should be given to adding these chemically reduced and organic forms of N r to the list of C ri teria P ol lutants.

Recommendation 8a: EPA should reexamine the criteria pollutant “oxides of nitrogen” and the indicator species NO 2 and consider adding chemically reactive nitrogen as a criteria pollutant, and NH x and NO y as indicators to supplement the NO 2 National Ambient Air Quality Standard.

Recommendation 8b: Monitoring of NH x and NO y should begin as soon as possible to supplement the existing network of NO 2 compliance monitors.

Recommendation 8c: EPA should pursue the longerterm goal of monitoring individual components of Nr, such as NO 2 (with specificity), NO and PAN, and HNO 3 , 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 dir ectly both at theCASTNET sites and in nearby locations with nonuniform surfaces such as forest edges.

• Improved measur ements and models of convective venting of the planetary boundary layer and of long range transport.

• Improved analytical techniques and observations

of atmospheric or ganic N compounds in vapor, particulate, and aqueous phases.

• Increased quality and spatial coverage of

measur ements of the NH 3 flux to the atmosphere from major sources especially agricultural practices.

• Improved measur ement techniques for, and numerical models of NO y and NH x species (especially with regard to chemical transformations, surfacedeposition 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 N r onto terrestrial ecosystems comes primarily from atmospheric deposition, synthetic fertilizer and B NF i n managed and unmanaged ecosystems ( T ab le 1). Although N r de posited from the atmosphere to terrestrial systems is formed inadvertently during fossil fuel combustion and from volatilization of NH 3 from agricultural activities, it serves to provide nutrients, along with biological N fi xation and synthetic fertilizer, for food, feed, and fiber production in the agricultural sector. F or ests and grasslands use N r for growth. H om e gardens, parks and recreational areas utilize N r within the urban landscape. Approximately 32 T g of new N r reached the land of the 48 contiguous states in 2002 ( T ab le 1). An additional ~0.2 T g of N w as imported mainly as food and drink products ( F A O
, 2007). An additional ~12 T g of N r

wa s recycled back to terrestrial and aquatic systems in livestock excreta (~6 T g N ), human excreta (~2 T g N ),

a nd crop residue from the previous year’s production (~4 T g N ; U.S. EP A, 2007e). O f this N
, ~ 1.3 T g (~1.2 from livestock manure and <0.1 from sewage sludge) was used as fertilizer for crop production (U.S. EP A, 2007e). More detailed information describing sources and cycling of reactive nitrogen input in terrestrial systems is provided in Appendix B. Finding 9: T otal N budgets within all terrestrial systems are highly uncertain. T he relative amount of N losses from terrestrial systems ascribed to leaching, runof f, 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, N r flows from the atmosphere and terrestrial systems into aquatic systems. Aquatic systems include groundwater, wetlands, streams and rivers, lakes and the coastal marine environment. N r is deposited directly into surface aquatic systems from the atmosphere (direct deposition) and N r that is not either stored or removed as products on terrestrial systems eventually moves into aquatic systems (indirect deposition).

T he area of an airshed generally greatly exceeds that of a watershed for a specific estuary or coastal region. F or example, the airshed of the Baltic Sea includes much of western and central E urope (Asman, 1994; H ov et al., 1994), while the airsheds of the two lar gest U.S. estuarine ecosystems, the C hesapeake Bay and Albemarle- P amlico Sound, are 15 to over 30 times the size of their watersheds ( D ennis, 1997). T hus, the airshed of one region may impact the watershed and receiving waters of another , making eutrophication a regional-scale management issue (Galloway and C owling, 2002; P aerl et al., 2002). F urthermore, atmospheric N inputs do not stop at coastal mar gins. Along the N orth American Atlantic continental shelf, atmospheric N inputs more than match riverine inputs (Jaworski et al., 1997; P aerl et al., 2002), underscoring the fact that N -driven marine eutrophication may require regional or even global solutions. E ven in truly oceanic locations (e.g., Bermuda), N orth American continental atmospheric N emissions (reduced and oxidized N ) are commonly detected and significant (Luke and D ickerson, 1987; P rospero et al., 1996). Likewise, islands in the N orth P acific receive N deposition originating in Asia ( P rospero et al., 1989).

R iverine and atmospheric “new” N r inputs in the N or th Atlantic O ce an basin are at least equal and may exceed “new” N r inputs from biological N 2 fixation ( H ow arth et al. 1996; P ae rl and Whitall, 1999; P ae rl et al. 2002). D uc e et al. (2008) estimate that up to onethird of the oceans’ external N r supply enters through atmospheric deposition. T hi s deposition leads to an estimated ~ 3% of new marine biological production and increased oceanic N 2 O production. Schlesinger (2009)

e stimated that global atmospheric transport of N r from land to sea accounts for the movement of almost one third of the annual terrestrial N r formation. T he refore, 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.

O ne example of shifting N inputs is the proliferation of intensive livestock operations in coastal watersheds, which has led to lar ge increases and changes in chemical composition of nitrogenous compounds discharged to estuarine and coastal waters via runoff, groundwater, and atmospheric deposition ( P aerl, 1997; H owarth, 1998; Galloway and C owling, 2002). I n 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 ( NH 4 + , organic N ) relative to oxidized N ( NO 3 - ) (Galloway and C owling, 2002; H owarth et al., 2002). T hese increases, combined with increases in hypoxia and anoxia in receiving waters, are leading to more NH 4 +

- rich conditions, which will favor those algal groups best able to exploit this N form, including some harmful algal bloom ( H AB) taxa ( P aerl and Whitall, 1999; P aerl 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 N r

. D evelopment also destroys wetlands, leading to more NO 3 - -enriched conditions, potentially favoring plant taxa best able to exploit this N form.

A

recent evaluation of decadal-scale changes of NO 3 - concentrations in ground water supplies indicates35 that there has been a significant increase in nitrate concentrations in well water across the U.S. (Rupert, 2008). T his 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
, N ational Water Quality Assessment P rogram. I n a subset of wells, ground water rechar ge 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 T errestrial Environmental Systems According to the nitrogen cascade conceptualization, terrestrial environmental systems are compartmentalized into agriculture, populated, and vegetated systems. Annual input of N r is greatest in agricultural ecosystems (farmland, cropland, and grazed pastureland). Annual N r in

puts to agricultural ecosystems, using 2002 as the base year, include 9.8 T g from synthetic fertilizer, 7.7 T g from biological N fi xation in crops (mainly soybeans), and 1.3 T g from atmospheric deposition. N r 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 T g) . Annual input of N r into populated systems includes synthetic fertilizer application to urban turfgrass and recreational areas (~1.1 T g) , and atmospheric deposition 0.2 T g.
Much of the annual N r input into these terrestrial systems passes through, and is transferred within, terrestrial systems or atmosphere via NH 3 , NO x or N 2 O
, or aquatic environmental systems via NO 3 - and organic N leaching and runoff or NH x and NO y deposition.

T he lar gest single reservoir of total N in the terrestrial environmental system is soil or ganic matter (S O M). Approximately 52,000 T g C and 4,300 T g 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). F or comparison, the total above ground biomass of U.S. forests of these states contains ~ 15,300 T g of C and ~ 59 T g N (estimated using a C / N ratio of 261), and 15,500 T g of soil or ganic matter carbon (S O M- C ), and 1,290 T g total N (estimated using a C / N ratio of 12) (U.S. EP A, 2007e). Most of this soil organic matter nitrogen (S O M- N ) is bound within complex or ganic molecules that remain in the soil for tens to thousands of years. A small fraction of this S O M is mineralized, converted to carbon dioxide and N r annually

. T he total N contained within above and below ground compartments isn’ t really of concern. What is of interest in addressing issues of N r , is the change in N stored within the compartments of terrestrial systems. T he pertinent question is whether N is being retained or released from long-term storage. T he C ommittee evaluated estimates of annual change of N storage within important components of terrestrial systems. T he C ommittee used C arbon stock information in C hapter 7 of the EP A I nventory of U.S. Greenhouse Gas E missions and Sinks 1990-2005 (U.S. EP A, 2007e) to estimate N storage in U.S. terrestrial systems. N itrogen 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 N r that moves from one part of the U.S. to another is the movement of N r 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. EP A SAB, 2007). T his 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 km 2 (Rabalais et al. 1996, 1999; Rabalais and T urner, 2001). T he 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. T he waters that dischar ge to the Gulf of Mexico originate in the watersheds of the Mississippi, O hio, and Missouri Rivers (collectively described here as the Mississippi River Basin). With a total watershed of 3 million km 2 , 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. EP A SAB, 2007).

T he D ecember , 2007 report, Hypoxia in the Northern Gulf of Mexico: An update by the EPA Science Advisory Board (U.S. EP A SAB, 2007) determined that, “ T o reduce the size of the hypoxic zone and improve water quality in the Basin, the SAB P anel recommends a dual nutrient strategy tar geting 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. F or both nutrients, incremental annual reductions will be needed to achieve the 45% reduction goals over the long run. F or 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 C roplands within the contiguous 48 states occupy ~ 368 million acres (149 million ha) (19%) of the 785 million ha of total land area. I n 2002, 126 million ha of this cropland were cultivated (US D A N R C S, 2007). C roplands are generally found on well drained mineral soils (or ganic C content 1-6% in the top 30 cm). Small areas of drained or ganic soils are cultivated (organic C content of 10-20%) in mainly F lorida, Michigan and Minnesota ( EP A, 2007e). O r ganic soils lost ~0.69 T g of N r in 2002 while mineral soils accumulated ~1.5 T g of N r ( T

able 10). Much of the accumulation of S O M- C was due to the use of conservation tillage and high yielding crop varieties (U.S. EP A, 2007e). Losses of N r from or ganic soils are due to mineralization of S O M and release of N r input. I n cultivated soils annual input of new N r is approximately 9.7 T g from fertilizer N , 1.1 T g from livestock manure (recycled N ), ~7.7 T g from biological N fixation, and 1.2 T g from atmospheric deposition. Assuming that loss of fertilizer N from the small area of

or ganic 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. EP A N at ional Greenhouse I nv entory (U.S. EP A, 2007g) the net increase in soil C s tocks over the period from 1990 through 2005 was largely due to an increase in annual cropland enrolled in the C on servation Reserve P ro gram, 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). I t is clear that conversion of marginal crop land to C R P
st ores C a nd reduces erosion and nitrate leaching. Likewise, use of soil conservation tillage, as opposed to conventional tillage with a plow or disk, reduces erosion. H ow ever, the impact of conservation tillage on soil C s torage and N r losses due to leaching or denitrification are much less certain. F or example, although the EP A estimates shown in T ab le 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 ( V er

ma et al., 2005; Baker et al. 2007; Blanco- C an qui and Lal, 2008). T he refore, the estimates of soil C a nd N s torage in mineral soils in T ab le 10 that were derived from U.S. EP A, (2007b) need to be reconsidered. T he se new studies and that of D av id et al. (2009) suggest that organic C c onservation 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 V

er

ma et al. (2005) also show that long-term, continuous gas exchange measurements have not detected C g ain due to no-till practices. T he y concluded that although there are other good reasons to use no-till, evidence that it promotes C s equestration is not compelling. T he se findings highlight the need for appropriate assessment of ecosystem N s torage in order to confirm or disprove this C om mittee’s conclusion that only a small part of annual N r input is stored in agricultural lands, forests, and grasslands.
Populated systems – urban lands P op ulated or “developed land” (developed land is the terminology used by the U.S D ep artment of Agriculture’s N at ural Resources C on servation Service [ N R C S] ) occupied ~106 million acres (42.9 million ha) of the U.S. land area in 2002. T hi s equates to approximately 5.5% of the U.S. land area (US D A N R C S, 2007). T he EP A I nv entory of U.S. Greenhouse Gas E mi ssions and Sinks (U.S. E nv ironmental P ro tection Agency, 2007e) indicates that urban areas cover over 4.4% of the land area with tree canopy covering 27.1% of the urban area. T he tree-covered area constitutes approximately 3% of total tree cover in the continental U.S. I f the N R C S 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. I n both urban forests and turf grass, N r st orage is dependent upon the age of the trees or turf. I n yo ung (pre-steady state) systems N i s being accumulated, while at steady state no net change occurs. Some areas may be degrading and actually losing biomass and returning N
t o the environment. EP A does not estimate carbon changes in turf grass, but does estimate changes in carbon storage in urban forests (U.S. EP A, 2007e). Urban trees sequestered an estimated net 22 T g of carbon and 0.12 T g of N i n 2002 (assuming that urban trees are mainly hardwood having

a C / N r atio of 186) (U.S. EP A, 2007e). Annual fertilizer N i nput into the urban landscape is approximately 10% of total fertilizer N c onsumption in the U.S. (U.S. EP A, 2 007e), or ~1 T g of N i n 2002. Another 0.2-1.0 T g N i s deposited from the atmosphere. I n some areas deposition can be disproportionately high due to locally high NO y concentrations. Storage of ~0.12 T g N i n urban forests constituted approximately 3% of N r input annually. Vegetated systems – forests and grasslands F or ests cover approximately 164 million ha, ~21% of the land area of the contiguous 48 states (US D A N R C S, 2007). T he forest carbon stocks analysis by EP A (U.S EP A, 2007e) is based on state surveys that are conducted every 1 to 10 years. Annual averages are applied to years between surveys. T o determine changes in C r elated to the rate of tree growth, Birdsey (1992) estimated that there were 52,500 T g of C a bove and below ground in U.S. forests; soil contained 59% of total C
, 9% was in litter, and 5% in tree roots. T he EP A estimate for 2002 was 43,600 T g of C

. T o estimate N s torage based on EP A data (U.S. EP A, 2007e), the C om mittee has assumed that forests are 85% softwood and 15% hardwood with an average C / N r atio of 261. T he se estimates indicate that forests and forest products stored ~0.43 T g of N i n 2002 ( T ab le 10).37 Grasslands, including rangelands and pasturelands, occupy approximately 213 million ha (27.1%) of the contiguous 48-state land area. T he N R C S divides these grasslands into pastureland (48.2 million ha) and rangeland (164 million ha). P astureland 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. C hanges in the N status of grasslands are dependent upon changes in soil or ganic 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 C entury biogeochemical model and data used were based upon the N R C S/ N ational Resources I nventory ( N R I ) survey (U.S. EP A, 2007e). C hanges in soil N content were estimated using a C / N ratio of 12. N itrogen input into rangelands is generally only from atmospheric deposition, which contributes 1.9 T g N each year to range production. Rangeland tends to be in relatively remote areas where atmospheric N r deposition is low

. C ollectively , forests and grasslands stored ~0.74 T g 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. F orest soils

T

able 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 Al though our estimate of N s torage in terrestrial systems is highly uncertain, annual N s torage is likely relatively small. Approximately 1.5 T g N w as stored in terrestrial systems of the contiguous 48 states in 2001 ( T

ab le 10). Soils were the largest reservoir with croplands (0.8) and grasslands (0.3) sequestering most of the N

. An nual storage in agricultural, grassland and forest soil and in forest biomass is approximately 6 to 10% of annual N r input. E st imated total N r input from synthetic fertilizer, biological N fi xation, and atmospheric deposition into terrestrial systems within the contiguous 48 states in 2002 was ~32 T g. All of the input and outflow numbers are highly uncertain, but N l oss 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 i nput while apparent loss through denitrification dominates the budget. T hi s is discussed in Section 2.3.3 of this report and in a recent global N r 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 I n considering N r transfers and transformations in and between the environmental systems of the nitrogen cascade, the C ommittee has encountered a number of areas where quantities or flows of N r are highly uncertain. All of these areas need attention from EP A 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 T otal 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. T hese gases include NO x, N 2 O and N 2.

n T he amount of N r transferred to each environmental

system as dry deposition needs to be quantified and monitored.

n T he fraction of NO y in the form of or gano-nitrates and

other organic nitrogen species is poorly quantified, but may play a major role in air quality and N r 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 T he annual change in N storage in soils (agricultural,

forest, grassland and urban areas) needs to be quantified in conjunction with the change in carbon. T hese areas of high uncertainty are highlighted because very little information exists in some of the areas. I n other areas, such as denitrification and the relative release of N 2 O from soils and aquatic systems, the sparse data are highly variable and this makes developing meaningful guidelines for control difficult. An analysis of N r input and fate in 16 watersheds in the U.S. is provided as an example to illustrate how the inputs and fate of N r can be evaluated for a lar ge watershed. Input and fate of Nr in 16 watersheds in the northeast United States T here 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

. D etermining a national N budget is a priority research

area. As there are no national data available, an example analysis of N r 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 N r for a lar ge watershed ( V an Breemen et al., 2002).

T he watersheds in this study encompass a range of climatic variability , Maine to V irginia. T he watersheds are a major drainage to the coast of the N orth Atlantic O cean. 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). T hey compared inputs with N losses from the system in riverine export. As a part of the same study, V

an Breemen et al. (2002) analyzed the fate of N inputs to these watersheds and developed budgets for each watershed. T he 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. T he N r input into the watersheds (using weighted averages for all 16 watersheds) was 3,420 kg per square km per year

. F igure 14 shows the N r sources and the estimated fate of this N r as a per cent of the weighted average N r input.

V

an Breemen et al. (2002) indicate that N r inputs and storages and losses were well correlated (R 2 = 0.98). D enitrification in landscape soils is the most uncertain estimate because rates are calculated by dif ference between total inputs and outputs so they accumulate errors from all estimates. T he denitrification loss term may also reflect the change in N storage in groundwater

. T he 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. I ncreasing storage of N r on land implies that drainage and denitrification exports of N r are likely to increase when a new steady-state condition is reached.

T he se data suggest that N r research needs to focus on understanding the “denitrification” loss term in this analysis. T he losses occur in the terrestrial landscape, before N r39 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 N r in watersheds to decrease nitrate movement into aquatic systems as well as to limit N 2 O emissions to the atmosphere.

T he V an 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. T he remaining one-third (1 1% of total N input) was considered to have been denitrified in the rivers. T hese examples also demonstrate that N r 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 N r enrichment problems. Aquatic and terrestrial systems process this N r and return other N r gases ( NH 3 , NO x and N 2 O to the atmosphere). N r from terrestrial systems impacts both the atmosphere and aquatic systems through emission of NH 3 , NO x , N 2 O and leaching and runoff of NO 3 -

. Finding 10: D enitrification of N r in terrestrial and aquatic systems is one of the most uncertain parts of the nitrogen cycle. D enitrification 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 5% Soil Storage

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.4041 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 or ganisms.

• Algal blooms eventually die off and consume

dissolved oxygen from the water column which can lead to die of f 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 or ganisms 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. Fi nding 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 Ecosystems42 

### 3.2. Impacts of Airborne or Atmospherically Deposited Nr on Human Health and Ecosystems Six major atmospheric effects are associated with increased NO x and NH 3 emissions, and two with N 2 O emissions (Galloway et al., 2003). For NO x and NH 3 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 NH 3 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, NH 3 , NO y , and

organic forms of Nr that can contribute to ecosystem acidification, fertilization, and eutrophication

For N 2 O, the effects are the greenhouse effect in the troposphere and O 3 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 ef fect.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 NO x , mg/L total NH x , or kg/ha of NO 3 - ). 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 damage44 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 SO 2 and NO x 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 negative46 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 NO x 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 NO x 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 adoptwater 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/NO x 14 $4,800

Industrial/NO x 15 $22,000

Mobile sources/ NO x 16 $14,000

Non-agricultural/NH 3 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 N 2 O 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,b48 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 O 2 , 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 forimproving 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 NO xemissions) 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.

• Ev aluate 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.

• De termine 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 r egulatory 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 ener gy use and greenhouse gas emissions, including nitrous oxide.

15c.

• Set Nr management goals on a regional/local basis,

as appr opriate, to ensure most effective use of limited management dollars.

• Fully consider “green” management practices such as

low- impact development and conservation measur es 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.

• Re search 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.

• Co nstruct a decision framework to assess and determine

implementation actions consistent with management goals.

15e.

• In cooperation with the Departments of Agricultur e 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 O 2 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 other50 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, NO 2 , ozone, SO 2 , and PM. These have been determined to endanger public health or welfare. The CAA as currently written requires areview of the scientific criteria for these standards at fiveyear intervals. Although NO 2 is the only Nr compound specified as a criteria pollutant, NH x and NO y 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 NO x 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 NO x emissions to below 1 g/mi (0.14 g N per km) until 1981. Few locales violate the standards for NO 2 , 18 but the secondary effects of several of these gases also pose health and welfare concerns. If a city had an annual average NO 2 level

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



### Pollutant Federal Primary Standard (NAAQS)
Ozone (O 3 ) 1-hr average 8-hr average 0.12 ppmv 0.08 ppmv

Nitrogen Dioxide (NO 2 ) 1-hr average Annual average 100 ppb 0.053 ppmv (100 μg/m 3 )

Particulate Matter, coarse (PM 10 ) Diameter ≤ 10 μm, 24-hr average Annual average

Particulate Matter, fine (PM 2.5 ) Diameter ≤ 2.5 μm, 24-hr average Annual average 150 μg/m 3 50 μg/m

35 μg/m 3 15 μg/m 3

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 NO 2 [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 NO 2 , it would risk severe photochemical smog – the summertime efficiency for ozone production ranges from 4 to 10 ppb O 3 per ppb NO x . As previously discussed, the focus on compliance monitoring for NO 2 ignores the other, equally important members of the NO y family such as HNO 3 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 NO x monitors to NO y monitors with a detection limit of 0.1 ppb would still demonstrate compliance with the NO 2 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 NO x , but not NH 3 or N 2 O. 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 SO 2 and 2 million tons of NO x 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, PM 2.5 , 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 NO x 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 NO 2 , which may not exceed 0.053 ppm (100 μg/m 3 ) 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 NO 2 concentrations well below 0.053 ppm lead to concentrations of secondary pollutants well above acceptable levels (i.e., PM 2.5 and O 3 ). The NO 2 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 NO 2 concentrations below 0.020 ppm (U.S.EPA, 2010a). The NO 2 concentration required to achieve the current 15 μg/m 3 PM 2.5 standard is probably also below the 100 μg/m 3 standard for NO 2 because of the role of NO 2 in secondary particulate formation. States in the eastern U.S. are considering substantial additional NO x 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 NO x emissions for point sources by 65%, (2) reducing NO x emissions for on-road sources by 75 percent, (3) reducing NO x 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 NO x 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 PM 2.5 and O 3 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 NO y and NH x .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 (NO y ) and chemically reduced (NH x ) 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 NO y and NH x 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 (NO x ) and Sulfur Oxides (SO x )” 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 not53 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 CO 2 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 NH 3 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 NH 3 volatilization as a means to remove N and decrease the N in the manure when land applied. Only recently has the resulting increase in NH 3 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 PM 2.5 concentrations. In this rulemaking, EPA states that “ammonia reductions may be effective and appropriate for reducing PM 2.5 concentrations in selected locations, but in other locations such reductions may lead to minimal reductions in PM 2.5 concentrations and increased atmospheric acidity.” Ammonia is a substantial component of PM 2.5 in most polluted areas of the United States at most times. While it is true that reducing NH 3 emissions might increase the acidity of aerosols and precipitation, the net effect of NH 3 on aquatic and terrestrial ecosystems is to increase acidity. After being deposited onto the earth’s surface, NH 4 + is under most circumstances quickly nitrified, increasing the acidity of soils and waters. The Committee is unaware of any evidence that NH 3 reduces the toxicity of atmospheric aerosols or that high concentrations of NH 3 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 PM 2.5 correlates with morbidity and mortality, and NH 3 contributes to PM 2.5 . The visibility degradation and other adverse effects associated with PM 2.5 are related to aerosol surface area or mass where NH 4 + certainly plays a role.

Recommendation 17: The Committee recommends that the EPA presumption that NH 3 is not a PM 2.5 precursor should be reversed and states should be encouraged to address NH 3 as a harmful PM 2.5 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 NH 4 + ion is the initial mineral N product formed during organic matter mineralization and most of the fertilizer used worldwide is NH 4 + based (e.g., urea, ammonium sulfate) (FAO, 2007), the suite54 of microbiological reactions that result in the release of gaseous N products need to be considered.
Nitrification is the oxidation of NH 4 + ion to NO 3 - (Figure 17). Most commonly, nitrification is a chemolithotropic process consisting of the conversion of ammonium to nitrite, which is then converted to NO 3 - 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 NH 4 + 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 NO 3 - and nitrite to produce NO, N 2 O, and N 2 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 NH 4 +

. Nitrous oxide reductase appears to be more sensitive to oxygen than either NO 3 - or nitrite reductase. Therefore N 2 production predominates in more anoxic sites and N 2 O production may be greater in more aerobic conditions. However, the ratio of N 2 to N 2 O emitted may also be affected by high NO 3 - concentrations and associated higher levels of electrical conductivity and osmotic stress and soil pH (low pH favors N 2 O 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 NO 3 - remediation at the expense of increasing input of N 2 O to the atmosphere. An example of such a situation involves NO 3 - 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 N 2 is produced. If, however, the process is incomplete, and NO and N 2 O 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 N 2 O is a GHG that has an atmospheric life time of approximately 100 years and a radiative forcing of approximately 300 times that of CO 2 on a hundred-year time frame (IPCC, 2001). N 2 O is also a major source of NO in the stratosphere and depletes stratospheric ozone (Crutzen, 1981). If more of the NO 3 - denitrified is converted to N 2 O in wetlands than upstream or downstream, the environmental cost may be high. Hernandez and Mitsch (2007) found that permanently flooded wetlands had lower N 2 O/N 2 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 matterto nitrate through the nitrification process. As nitratecontaining effluent from sewage treatment flows into aquatic systems the nitrate may be denitrified, resulting in N 2 O 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 N 2 O-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 N 2 O 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 nitr ogen;

n Implementing a resear ch program that addresses the impacts of tradeoffs associated with management strategies for carbon, reactive nitrogen, phosphorus, and other contaminants of concern;

n Implementing a resear ch 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 CO 2 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 Dischar ge 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 N 2 O and O 3 are greenhouse gases.

n Soil Nr chemistry and emissions of N 2 O, NH 3 , 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 NO x 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 r eport for the year 2050.

2. Down-scale statistics or nest regional climate models within each of these global scenarios to generate
meteor ological and chemical fields (e.g., temperature, relative humidity, winds, precipitation, CO 2 ) 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 curr ent Nr and emissions and application rates.

4. Rerun models with decreased Nr emissions/application to evaluate strategies for contr olling 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 NO 3

'96 '89 '88 '95 '94 '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/yrProbability Lognormal Distribution
Total NO 3 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 Kalkhof f 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 dischar ge, as represented by data for the years 1993 and 1998. Figure 20 shows that the interannual variation in nitrate dischar ged 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 dischar ge 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.5859 

# 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 NO x 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 dischar ge 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-control 19Policy 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 atleast 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 costs61 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 Control 20 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. Wa ter 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. T wo 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 T ransferable 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.

Po licy 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 thepotential 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, M arket 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 N 2 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 NO x 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 dif ference 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, 200765 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, 200766 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 needed67 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 NO 3 - 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., NH 3 , NO x , N 2 O, NO 3 - ) 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 NO 3 - 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 NO 3 - converted to N 2 (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 NO 3 - is converted mainly to N 2 that is emitted to the atmosphere (Clement et al., 2002). As discussed in Section 4.7, the potential for the formation of N 2 O 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 NO 3 - 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 NO 3 - -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 (2million 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 amajor, 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: NH 3 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 NO x , and is available for additional N 2 O production and movement into aquatic systems as NH 4 + and NO 3 -

. 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 N 2 . The fraction of manure N that can be and is converted to N 2 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 NH 3 , NO, N 2 , and N 2 O 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 NH 3 emissions after a permeable cover was installed. Miner et al. (2003) reported that a polyethylene cover can reduce NH 3 emissions by ~80%, but it is not clear what fraction of that N was converted to N 2 . Harper et al. (2000) reported that in a well-managed swine lagoon denitrification N 2 losses can be equivalent to N lost as NH 3 , in other words about 50% efficiency. Kermarrec et al. (1998) reported that sawdust litter helped reduce NH 3 emissions from pig manure with 44-74% of manure71 N converted to N 2 , but greater than 10% of the manure N was released as N 2 O. Sommer (1997) found that NH 3 was emitted from cattle and pig slurry tanks at the rate of 3.3 kg N m 2 yr -1 until covered with straw. After straw application NH 3 emissions were below detection limit. Mahimairaja et al. (1994) reported that NH 3 volatilization was reduced by 90-95% under anaerobic conditions. Section 2.2.4 contains a discussion of best management practices to minimize NH 3 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: NO x As previously discussed, a major contributor to Nr in the atmosphere is fossil fuel combustion. During the combustion process NO x (NO x = NO + NO 2 ) are released to the atmosphere. Globally the production of NO x 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 NO x 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 NO x 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 Th ermal NO x where N 2 and O 2 combine at high

temperatures (thermal pathway dominates at temperatures greater than approximately 1500 C) to form NO through the Zeldovich mechanism

n Fu el NO x where nitrogen from a fuel (e.g., coal and

biofuels) is released as some intermediate and then combines with O 2 to form NO

n Pr ompt NO x where N 2 reacts with hydrocarbon radicals

in flames, forming various compounds including hydrogen cyanide and other cyano radicals. These in turn form NO x . Contributions of prompt NO x are usually low as compared to fuel NO x . There are several ways to control emissions of NO x . 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/O 2 reaction. Temperature can be controlled by using a fuel-rich mixture versus fuel lean. In this casethe reactions take place at lower temperatures. Fuelrich mixtures also reduce the amount of O 2 available for reaction and there are changes in the chemical mechanisms which limit the oxidation of N 2 . If fuellean mixtures are used for temperature control, while the temperature is lower there is a significant amount of O 2 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-NO x burners operate under the principle of internally staging the combustion. To reduce fuel NO x , air and fuel staging are used to reduce the peak temperature where air and fuel are admitted in separate locations.
Chemical reduction of NO x 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 NH 3 is injected in a controlled temperature zone to allow the reduction of NO x . SCR is also an add-on technology where the flue gas must pass through a catalyst bed to allow reaction between ammonia and NO x . Care must be taken with both technologies to avoid NH 3 slip. Fuel reburning requires the injection of a fuel to create a zone where NO x is reduced to N 2 . Low NO x burners may also use an internal fuel reburning to reduce the NO x . 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 NO x versus fuel NO x 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 NO x burners, while spark ignition engines utilize a three-way catalyst which requires less than 0.5% O 2 . In this case, additional NO x is reduced by utilizing unburned fuel as a reagent over the catalyst for chemical reduction of NO x . It should be noted, however, that a side reaction for the three-way catalyst system produces ammonia. NO x emissions can be reduced in diesel engines by delaying theinjection 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 NO x removal from the American light-duty fleet (Bishop and Stedman, 2008).7273 

# 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 the74 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 and75 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, theCommittee 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) improvedremoval 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 forassessing 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 loadings76 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 NO x

emissions from mobile and stationary sources The Clean Air Act (1970) and its Amendments (1990) have resulted in NO x 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. NO x 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 NO x 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 NO x 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 NO x emissions are implemented only in the ozone season (May to September) (U.S. EPA 2009c). To protect welfare and avoid adverse effects on ecosystems, NO x 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 NO x 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% NO x 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 NO x 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 the77 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 N 2 O 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 N 2 O 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 N 2 O 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 NO x 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 NH 3 emissions that have been developed since 1990 will further decrease ammonia and other gases and odor emissions. For example, sawdust litter helps decrease NH 3 emissions from pig manure with 44-74% of manure N converted to N 2 . 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 NH 3 emissions after a permeable cover was installed (e.g., a polyethylene cover decreased NH 3 emissions by ~80%). A well managed swine lagoon can denitrify approximately 50% of the excreted N to N 2 . 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 NH 3 emissions from swine and poultry production, a moderate reduction of 50% from 1990 NH 3 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 NH 3 emissions, a more modest and reachable goal of decreasing NH 3 emissions by 10% through improvements in animal diet and manure management is proposed (Table 17). Management Goal 3

. The Committee estimates that livestock-derived NH 3 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 PM 2.5 by ~0.3 μg/m 3 (2.5%), and improve health of ecosystems by achieving progress towards critical load recommendations. Additionally we estimate that NH 3 emissions derived from fertilizer applications can be decreased by 20% (a decrease78 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 NH 3

emissions from livestock manure in the United States



### NH 3 

### Source % of Total NH 3 

### Tg NH 3 

### -N/yr emitted Estimated Decrease of NH 3 

### % 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 0.4579 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 af fecting 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 dif ferent 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. I n 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. Ex amine 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 tradeof fs across impact categories involving the cycling of other elements, particularly carbon and phosphorus. (Finding #18)

n Th e 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 of81 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 Th e 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 Committeerecommends 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 Sc ientific 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, NO 2 , is often a small component of NO y , the total of oxidized nitrogen in the atmosphere. The current NAAQS for NO 2 , as an indicator of the criteria pollutant “oxides of nitrogen,” is inadequate to protect health and welfare. NO y should be considered seriously as a supplement or replacement for the NO 2 standard and in monitoring. Atmospheric emissions and concentrations of Nr from agricultural practices (primarily in the form of NH 3 ) have not been well monitored, but NH 4 + ion concentration and wet deposition (as determined by NADP and NTN) appear to be increasing, suggesting that NH 3 emissions are increasing. Both wet and dry deposition contribute substantially to NH x 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

• Ph ysical, chemical or biological “dead ending” or

storage within landscape compartments where it82 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 runof f 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 Cu rrent 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 PM 2.5 concentrations. In this rulemaking ( Federal Register 72(79): 20586-20667), EPA has stated that “ammonia reductions may be effective and appropriate for reducing PM 2.5 concentrations in selected locations, but in other locations such reductions may lead to minimal reductions in PM 2.5 concentrations and increased atmospheric acidity.” Ammonia is a substantial component of PM 2.5 in most polluted areas of the U.S. at most times. While it is true that reducing NH 3 emissions might increase the acidity of aerosols and precipitation, the net effect of NH 3 on aquatic and terrestrial ecosystems is to increase acidity. After being deposited onto the earth’s surface, NH 4 + is, under most circumstances, quickly nitrified, increasing the acidity of soils and waters. The Committee is unaware of any evidence that NH 3 reduces the toxicity of atmospheric aerosols or that high concentrations of NH 3 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 PM 2.5 correlates with morbidity and mortality, and NH 3 contributes to PM 2.5 . The visibility, degradation, and other adverse effects associated with PM 2.5 are related to aerosol surface area or mass where NH 4 + 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, NO 2 , and consider adding chemically reactive nitrogen as a criteria pollutant and NH x and NO y as indicators to supplement the NO 2 National Ambient Air Quality Standards. (Recommendation #8a)

n The Committee recommends that monitoring of NH x

and NO y begin as soon as possible to supplement the existing network of NO 2 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 measures83 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 ef fects 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 NH 3 is not a PM 2.5 precursor should be reversed and states should be encouraged to address NH 3 as a harmful PM 2.5 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 Ni trous 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, N 2 O, reduced sulfur compounds, VOCs, and NH 3 . 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 (SO 4 2- ), NO 3 - , and NH 4 + 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-making84 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)

• Ef forts 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., NO 3 - and NH 4 + ) 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 Th rough 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 NO 2 (with specificity), NO, PAN, and HNO 3 , 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 NH 3 flux to the atmosphere from major sources especially agricultural practices

• Im proved measurement techniques for, and numerical models of, NO y and NH x species (especially with regard to chemical transformations, surfacedeposition 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 Th e 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, the85 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, CO 2 ) 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 dif ferently

• 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 cascade86 (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 that87 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.8889 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, NO 3 - , NH 4 + , 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 NH 4 + (reduced N), and NO 3 - (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 NO 3 - wet deposition, but not in NH 4 + deposition (Table A-1 and Figure A-2). Ammonium wet deposition shows a weak increase, although the correlation coefficient is small. As NO x 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 NO x emissions occurred as the result of the NO x 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 NO 3 - and NH 4 + 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 (NH 4 +

), oxidized (NO 3 - ), 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 (NO 3 - ) and reduced (NH 4 +

) nitrogen wet deposition from 1994 to 2006.
As emissions of NO x have decreased, the relative importance of NH x has increased (data from National Atmospheric Deposition Program, 2010).Atmospheric wet N deposition USA



### NADP/NTN deposition estimates90 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 – NO 3 - and NH 4 +

. Decreases in NH 4 + concentration and wet deposition are attributed to decreases in SO 4 2– concentrations, meaning that more of the reduced Nr remains in the gas phase. For the period 1965 to 2000, NO 3 - levels in bulk deposition correlate well with reported NO x 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 NO x 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 NH 4 + 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/yrNitrate deposition, kgN/ha/yr

Figure A-2: Trend in reported wet deposition of NH 4 + and NO 3 - 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 NH 4 + , NO 3 - , and total inorganic N deposition for the year 2007 showing spatial patterns of deposition Source: National Atmospheric Deposition Program, 2010.91 Concentrations of aerosol NH 4 + 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 NH 4 + will be accompanied by an increase in vapor phase NH 3 if SO 4 2– and NO 3 - concentrations decrease; see http://vista.cira.colostate.edu/improve/. This potentially misleading information highlights the need for measurements of speciated NH x (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 HNO 3 , particulate NO 3 - , and NH 4 + ; it does not include deposition of other oxidized species such as NO x and PAN, nor gas-phase reduced N species, most notably NH 3 . 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 HNO 3 and NO 3 - , but not other members of the NO y family – notably NO x . Dennis (U.S. EPA, 2007d) estimated that the unmeasured NO y 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 NH 3 indicate that dry deposition should account for 75% of wet NH 4 + 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 N 2 and N 2 O via gas phase reaction is provided in Appendix E. Warmer temperatures are conducive to release of NH 3 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 HNO 3 deposition velocity is 1.24 cm/s while that for particulate NO 3 - is 0.10 cm/s. In 2003 and 2004 substantial reductions in emissions from electricity-generating units (power plants) were implemented under the NO x 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 NO y 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 NH 4 + 0.41
Wet NH 4 + 2.54 Dry HNO 3 + NO 3 - 1.88
Wet NO 3 - 2.92 Total measured N Dep. 7.75 Est. dry other NO y 0.94 Est. dry NH 3 1.90
Est. total NO y 5.74
Est. total NH 3 + NH 4 + 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 HNO 3 , particulate NO 3 - , and NH 4 + ; unmonitored are other oxidized species such as NO x and PAN and gas-phase reduced N species most notably NH 3 (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 HNO 3 . 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 NH 3 ), 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 NO x emissions from marine vessels, and these amount to about 4% of the total NO x emissions in 2000. Calculated nitrogen deposition for the 48 contiguous states (Table A-3) was broadly consistent with direct measurements (Table A-2). CMAQ NO x 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, NH 3 emissions were derived from observed NH 4 + wet deposition (Gilliland et al., 2003; Mathur and Dennis, 2003; Gilliland et al., 2006). Model determinations of NH 3 therefore do not provide an independent source of information on NH 4 + deposition. The three-year CMAQ run gives an indication of the spatial pattern of deposition (Figures A-4). For NH x , wet and dry are equally important, but for NO y , 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 NO y deposition is 2.79 times the wet deposition and total NH x 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 NO x SIP-call, which resulted in substantial reductions in NO x 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 NO x is assumed to be primarily HNO 3 or aerosol NO 3 - ; organic N species are generally not modeled in detail. Because this analysis includes Alaska, a better estimate for NO x 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 NO y and NH x 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 NH 3 , NO x 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 NO y and NH x 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 NO x is exported in the lower to mid-troposphere. CTMs derived small export values – on the order of 30% of the total NO x 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 NO y emitted over the U.S. is exported. Comparison of models and measurements of oxidized N deposition Both ambient measurements and numerical models of NO y 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 NO x emissions are exported. Results from CMAQ runs described above indicate that, of the NO x 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 NO 2 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, 2007f94 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 NO 3 - , volatilized to the atmosphere as NH 3 or NO x or denitrified (see Section 4.8, Figure 18) to produce NO x , N 2 O and N 2 . 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 NH 3 volatilization and NO 3 – 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 NH 3 volatilization, nitrification/ denitrification, and NO 3 - 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 NO 3 - (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.9899 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 NO 3 – ) 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 (NO 3 - ) 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 Basin100 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 Communication101 “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.102103 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 NO 2 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 Loads104 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 NO x emissions 3 Convention on Longrange Transboundary Air Pollution

Clean Air for Europe Other photochemical oxidants Org. NO 3 , PAN concentration (atm) NO NO x emissions 5 Convention on Long-range Transboundary Air

Pollution et al. Fine particulate aerosol PM 10 , PM 2.5 concentration (atm) YES NO x , NH 3 em 1 Convention on Longrange Transboundary Air Pollution

Clean Air for Europe Direct toxicity of nitrite NO 2 - NO 2 - concentration YES NO x 2 World Health Organization

Convention on Longrange Transboundary Air Pollution

Clean Air for Europe
Nitrate contamination of drinking water NO 3 - concentration (aq.) YES NO 3 - leaching 2 EU Essential Facilities Doctrine

Depletion of stratospheric ozone NO x , N 2 O concentration/

flux (atm) NO NO x , N 2 O 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

NO 3 - (&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 (O 3 flux), AOT40** YES NO x 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) BiologicalOxygen Demand, NO 3 - conc (aq)

Critical loads YES

NO Runoff, Nr deposition 3 Water Framework Directive

Eutrophication of coastal ecosystems inducing hypoxia (incl. biodiversity) BOD, NO 3 - conc (aq)

Critical loads BOD, NO 3 - 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 (NH 3 , NO x ) 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 billion106 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., O 3 ,

PM YES NO x , NH 3 3 Convention on Long-range Transboundary Air Pollution
Effects on monuments and engineering materials PM 2.5 conc (atm) NO NO x , NH 3 4

Global climate warming induced by excess nitrogen N 2 O, conc/flux (atm) NO NO x , NH 3 1 United Nations Framework Convention on Climate Change Regional climate cooling induced by aerosol) PM 2.5 conc (atm) NO NO x , NH 3 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 N 2 and N 2 O via gas-phase reactions Atmospheric conversion of NO x and NH x to less reactive N 2 or N 2 O 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 NH 3 and NO x to N 2 and N 2 O, start with attack of NH 3 by OH:

NH 3 + OH· → NH 2 · + H 2 O (1) Several potentially interesting fates await the NH 2 radical:

NH 2 · + O 3 → NH, NHO, NO (2)

NH 2 · + NO 2 → N 2 O + H 2 O (3)

NH 2 · + NO → N 2 + H 2 O (4)
kO 3 = 1.9x10 -13 cm 3 s -1

kNO 2 = 1.8x10 -11 cm 3 s -1

kNO = 1.8x10 -11 cm 3 s -1 The first step, attack by OH, is slow. The rate constant for the Reaction 1 is 1.6 x 10 -13 cm 3 s -1 and the lifetime of NH 3 for a typical concentration of 106 OH cm -3 is about 70 days. In most areas of the world where concentrations of NH 3 are high, concentrations of sulfates are also high, and NH 3 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 NH 3 and NO x are emitted in large quantities, but SO 2 is not, and there gas-phase conversion can take place. NH 3 is usually removed by wet or dry deposition. Reaction 2 is relatively slow and oxidation of ammonia is in general an unimportant source of NO x . But Reactions 3 and 4 may be atmospherically noteworthy. As an upper limit to current N 2 O production, we can assume that each of these regions covers an area of 10 6 km 2 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 N 2 and/ or N 2 O would then be on the order of 0.1 Tg N, a minor but nontrivial contribution to denitrification and about 1% of the anthropogenic N 2 O production. If NH 3 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 N 2 O or N 2 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 NH 3 emissions and the atmospheric source of N 2 O will grow in importance. In the stratosphere, N 2 O photolysis leads to loss of Nr via

N 2 O + hμ → N 2 + O (5) While reaction with an electronically excited oxygen atom O(1D) leads to production of NO via

N 2 O + O( 1 D) → 2NO (6) Photolysis (Reaction 5) dominates, but a large enough fraction of the N 2 O reacts with O( 1 D) that this is the main source of NO x in the stratosphere. The fate of this oxidized nitrogen (NO y ) is transport back into the troposphere where it is removed by wet deposition. Downward transport of the odd N from the oxidation of N 2 O is a minor (~1%) source of NO y in the troposphere. Most of the N 2 O released into the atmosphere is eventually converted to N 2 – 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 N 2 and N 2 O 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 modelingapproach 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 Annexes108 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., 2008109 EPA informed the Committee that it is in the process of implementing a number of regulations to reduce NO x from a variety of mobile sources 22 . These include clean diesel regulations for trucks and buses and nonroad engines, as well as locomotives and smaller marine vessels. EPA first regulated NO x emissions from motor vehicles for the 1973 model year and since then has tightened these standards. EPA’s efforts to control NO x emissions from nonroad vehicles, locomotives, and commercial marine vessels started in the 1990s. NO x reductions for each rule were calculated by EPA based on inventories available at the times of the rules.

1. Li ght 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 NO x reductions.

2. EP A’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 NO x 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 NO x emission levels that are 95% below the pre-rule levels. EPA projects a 2.6 million ton reduction of NO x 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. Cl ean 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 NO x by about 740,000 tons. 4. Marine-Related NO x Reductions from 1999 to 2003. EPA completed three rulemakings with respect to the diesel marine sector that will reduce NO x emissions. These rules are now in effect and being phased-in. In 1999 (64 FR 73299, December 29, 1999), EPA promulgated NO x requirements for diesel enginesused 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 NO x 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 NO x from diesel engines used in commercial vessels by about 20%. EPA projects that on a nationwide basis, these four programs will reduce marine-related NO x 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 NO x 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 NO x emissions reductions of 80% from engines meeting these standards. EPA projects that in 2030, about 420,000 tons of NO x will be reduced from the locomotive engines, and 375,000 tons of NO x 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+NO x emissions from new engines’ exhaust, reduce evaporative emissions by 45%, and that together these programs will reduce NO x 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 NO x

22 The information in Appendix F was provided to the Integrated Nitrogen Committee by Mazrgaret Zawacki of the U.S. EP A 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 NO x emissions from ocean-going vessels by 80% and that in 2030, the coordinated strategy is expected to yield a reduction in NO x 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 NO x ) 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 NO x 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 NO x 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 NO 3 . It has been long established that excess NO 3 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 NO 3 in drinking water, EPA has established a maximum contaminant level goal (MCLG) that considers a lifetime exposure plus a margin of safety. For NO 3 in drinking water, the MCLG is 10 mg/L. Nitrite-N also has an established MCLG of 1 mg/L, and the combined NO 3 and NO 2 MCLG set by EPA is 10 mg/L. These same values are used to regulate NO 3 and NO 2 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 NO 3 -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 NO 3 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 (r 2 = 0.801). As illustrated in Figure G-2, surveys confirm that NO 3 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 NO 3 -N levels (median concentration of 3.4 mg/L). Urban shallow aquifers were less impacted (median concentration of 1.6 mg NO 3 -N/L) and deeper, major 

# Appendix G: Impacts of Reactive Nitrogen on Aquatic Systems112 aquifers, in general, had a median NO 3 -N concentration of 0.48 mg/L. However, NO 3 -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 NO 3 -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 NO 3 -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 NO 3 -N. DeSimone et al. (2009) suggested that redox could be a defining factor in some cases, and perhaps was the reason for low NO 3 -N concentrations in the Southeast in soils that promote denitrification, as well as higher NO 3 -N levels in other areas where aquifers were better oxygenated. Ammonia toxicity in freshwater systems The EPA and states have long regulated ammonia (NH 3 ) 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 (NH 4 + ), and research has further demonstrated the relationship between pH, temperature, and NH 3 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 NH 3 , 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 NH 4 + 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., 2009114 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, includingnoticeable “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, 1999.115P 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-N 2 - fixing cyanobacteria, such as Microcystis, are superior competitors for reduced N (Blomqvist et al., 1994), but even N 2 -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, andeutrophication 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 and116 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) characterizedthis 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 to117 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 (NH 3 and NH 4 + ion) and oxidized (NO 3 – ) 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 O 2 or118 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 & Plan119 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 O 2 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 O 2 criteria in both states. The revised TMDL will likely require more aggressive reductions of nitrogen to meet dissolved O 2 criteria and may formalize targets for upstream state contributions and atmospheric deposition.120121 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 NH 4 + builds up it is converted to nitrate by bacterial action, a process that releases hydrogen ions and contributes to soil acidification. The buildup of NO 3 ˉ enhances emissions of nitrous oxides from the soil and also encourages leaching of highly water-soluble NO 3 ˉ into streams or groundwater. As negatively charged NO 3 ˉ 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 muchless 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). 

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