National Acid Precipitation Assessment 

Program Report to Congress 2011:

An Integrated Assessment

2011
Photos courtesy of the following:

Cover – National Park Service

Executive Summary – Dani Newcomb

Introduction – U.S. Geological Survey

Chapter 1 – Jane Hawkey, IAN, UMCES

Chapter 2 – U.S. Geological Survey

Chapter 3 – Dani Newcomb

Chapter 4 – U.S. Geological Survey

References – National Park Service
EXECUTIVE OFFICE OF THE PRESIDENT 

NATIONAL SCIENCE AND TECHNOLOGY COUNCIL 

WASHINGTON, D.C. 20502 

December 28, 2011 

Members of Congress: 

I am pleased to transmit the National Acid Precipitation Assessment Program Report to Congress: 

An Integrated Assessment. The National Acid Precipitation Assessment Program (NAPAP) is a Federal 

interagency program that coordinates acid rain research and reports to Congress on the effects of acid 

precipitation on sensitive ecosystems. Specifically, this report presents the latest scientific information and 

analysis concerning the costs, benefits, and environmental effectiveness of the Acid Rain Program (ARP)—a 

bipartisan mandate under Title IV of the 1990 Clean Air Act Amendments signed into law by President 

George H. W. Bush to reduce sulfur dioxide (SO2) and nitrogen oxides (NOX) emissions from electric 

generating sources. Notably, the SO2 program includes the use of a creative emissions cap-and-trade 

program that combines the best of American science, government, and market-driven innovation. 

This NAPAP Report to Congress focuses on emission reductions from power plants, summarizes 

changes in deposition rates and environmental impacts, and evaluates the ecological effects expected to 

accompany future reductions in SO2 and NOX emissions. The report concludes that Title IV has successfully 

reduced emissions of SO2 and NOX from power generation (i.e., the sources covered by the ARP), including: 

SO2 emissions to 5.7 million tons in 2009, 64% lower than 1990 emissions and below the 2010 statutory cap 

of 8.95 million tons; and NOX emissions to 2 million tons in 2009, 67% lower than 1995 emissions and 

substantially exceeding the Title IV goal. The emission reductions achieved under the ARP have contributed 

to measurable improvements in air quality; decreases in acid deposition; the beginnings of recovery of acidsensitive lakes and streams in some areas; and improvements in visibility. 

This report also estimates the benefits and costs of complying with Title IV. The human health 

benefits of improved air quality are estimated to be in the range of $170 billion to $430 billion in 2010 alone. 

Substantial additional benefits result from improved visibility and improved ecological conditions. These 

benefits greatly exceed the cost of complying with Title IV, which are estimated at about $3 billion per 

year—less than half the initial estimates of compliance with Title IV’s tightening of the SO2 emission cap in 

the year 2000, which now applies to fully 96 percent of the Nation’s total electricity generation from fossil

fuels. The use of a cap-and-trade mechanism significantly lowered costs by giving utilities flexibility in 

achieving SO2 emissions reductions. 

Despite the environmental improvements reported here, research over the past few years indicates 

that recovery from the effects of acidification is not likely for many sensitive forests and aquatic ecosystems 

without additional decreases in acid deposition. Scientific publications and acid deposition models show that,

for these sensitive regions, the SO2 and NOX emission reductions achieved under Title IV from power plants 

are insufficient to achieve full recovery or to prevent further acidification.

Sincerely,

John P. Holdren 

Assistant to the President for Science and Technology 

Director, Office of Science and Technology Policy 
i

Acknowledgments

The National Acid Precipitation Assessment Program (NAPAP) Office of the 

Director would like to thank all those who wrote, edited, and reviewed this NAPAP 

Report to Congress (RTC). Special thanks are extended to the extramural peer 

reviewers, who, in addition to providing a technical review, were asked to consider 

the relationship of this report to public policy. Their comments and suggestions make 

this a significantly better report.

NAPAP Director and Report Lead Author

Douglas A. Burns, U.S. Geological Survey

Primary Authors and Contributors

Executive Summary and Introduction

 All contributing authors

Chapter 1 

 Clean Air Markets Division, U.S. Environmental Protection Agency

Chapter 2 

 Clean Air Markets Division, U.S. Environmental Protection Agency 

 Douglas A. Burns, U.S. Geological Survey

Chapter 3 

 Douglas A. Burns, U.S. Geological Survey 

 Mark E. Fenn, U.S. Forest Service 

 Jill S. Baron, U.S. Geological Survey

Chapter 4 

 Jason A. Lynch, U.S. Environmental Protection Agency 

 Bernard J. Cosby, University of Virginia
ii

Peer Reviewers

Dallas Burtraw, Resources for the Future 

Keith N. Eshleman, University of Maryland Appalachian Laboratory 

George M. Hidy, Envair/Aerochem 

Dean S. Jeffries, Canada Centre for Inland Waters, Environment Canad

Paul J. Miller, Northeast States for Coordinated Air Use Management 

William H. Schlesinger, Cary Institute of Ecosystem Studies 

John L. Stoddard, U.S. Environmental Protection Agency

About the National Acid Precipitation Assessment Program

NAPAP is a cooperative Federal program first authorized in 1980 to coordinate 

acid rain research and report the findings to Congress. The research, monitoring, 

and assessment efforts by NAPAP and others in the 1980s culminated in Title 

IV of the 1990 Clean Air Act Amendments (CAAA), also known as the Acid 

Deposition Control Program. In a bold new approach to environmental protection, 

Title IV includes a market-based program that provides economic incentives for 

controlling emissions of sulfur dioxide from electricity generating facilities. Title IX 

of the CAAA reauthorized NAPAP to conduct acid rain research and monitoring 

and to periodically assess the costs, benefits, and effectiveness of Title IV. The 

NAPAP member agencies are the U.S. Environmental Protection Agency, the U.S. 

Department of Energy, the U.S. Department of Agriculture, the U.S. Department 

of Interior, the National Aeronautics and Space Administration, and the National 

Oceanic and Atmospheric Administration. This report is the fourth published by 

NAPAP since 1990 assessing Title IV.

In 1997, NAPAP began to operate under the auspices of the Committee on 

Environment, Natural Resources, and Sustainability (CENRS), of the National 

Science and Technology Council (NSTC). NAPAP’s goal continues to be providing 

credible technical findings on acid deposition and its effects to inform the public 

decision-making process. To ensure that this goal is met, NAPAP coordinates its 

activities through the Air Quality Research Subcommittee of the CENRS.

 

 

 

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iii

Contents

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

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .1

Chapter 1 Acid Rain Program Elements and Implementation . . . . . . . . . . . . . . . . . .3

Overview of Emission Reductions ........................................ 3

Acid Rain Program Design .............................................. 4

SO2 Program................................................... 4

NOx Program .................................................. 6

Title IV Affected Sources ............................................... 7

SO2 Sources ................................................... 7

NOx Sources ................................................... 7

Compliance ......................................................... 7

SO2 Compliance................................................ 7

NOx Compliance ............................................... 8

2009 SO2 Allowance Market....................................... 8

Program Benefits and Costs ............................................ 12

Benefits ...................................................... 12

Costs........................................................ 16

Program Assessment Tools ............................................. 17

Emission Monitoring and Reporting ............................... 17

Air Quality, Deposition, and Ecological Monitoring .................... 17

Chapter 2 Results of the Acid Rain Program: Status and Trends of Emissions, 

Deposition, Air Quality, Surface Water, and Visibility, 1990 to 2009 . . . . . . . . . . .21

Emissions .......................................................... 21

SO2 Emissions................................................. 21

NOx Emissions ................................................ 23

Air Quality ......................................................... 25

SO2 and Sulfate................................................ 25

NOx and Nitrate............................................... 28

Acid Deposition ..................................................... 30

Sulfur Deposition .............................................. 31

Nitrogen Deposition (Nitrate and Ammonium) ....................... 32

Surface Water Monitoring ............................................. 33

Case Study—Critical Loads for Lakes and Streams ..................... 37

Visibility .......................................................... 42
iv

Chapter 3 Effects of Acid Deposition on Ecosystems: 

Advances in the State-of-the Science . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .45

Effects of Atmospheric Deposition on Aquatic and Terrestrial Ecosystems ......... 46

Ecosystem Recovery in Aquatic and Terrestrial Ecosystems of the 

Eastern United States ......................................... 46

Current Effects of Atmospheric Deposition on Aquatic and Terrestrial 

Ecosystems in the Rocky Mountain Region of the United States......... 52

Current Effects of Atmospheric Nitrogen Deposition on Terrestrial and 

Aquatic Ecosystems in the Far Western United States ................. 54

Effects of Atmospheric Nitrogen Deposition on Coastal Estuaries ............... 56

Emerging Issues at the Interface of Science and Policy in Managing 

the Effects of Acid Deposition on Ecosystems ............................ 57

Critical Loads ................................................. 57

Atmospheric Deposition and Ecosystem Services....................... 59

What are the Implications of Ongoing and Future Changes in Climate and the 

Carbon Cycle for the Effects of Acid Deposition on Ecosystems?.............. 61

Implications of Climatically Driven Changes in the Species 

Composition of Ecosystems .................................... 62

Climate Warming Will Affect Many Temperature-Sensitive 

Biogeochemical Processes in Ecosystems ........................... 62

Implications of Changes in Carbon Dioxide Concentrations 

in the Atmosphere............................................ 64

Changes in the Water Cycle are Key to Understanding 

Ecosystem Effects ............................................ 64

Nitrogen Availability has Important Implications for the Response 

of Ecosystems to Climate Change................................ 65

Dissolved Organic Carbon—Acid Deposition Interactions: 

A Case Study of Climate Change ................................ 66

Summary of Climate Change— Atmospheric Deposition Linkages......... 68

Multi-Pollutant Interactions in Ecosystems................................. 69

Ozone Interactions with Atmospheric Sulfur and Nitrogen Deposition...... 69

Mercury Interactions with Atmospheric Sulfur and Nitrogen Deposition .... 70

Chapter 4 Beyond Title IV – Ecological Impacts of Further Emission Reductions . .71

Recent Assessments................................................... 72

Analysis of the Environmental Impact of Further Air Emission Reductions ........ 73

Scenarios Analyzed ............................................. 74

Changes in Sulfur Deposition ..................................... 77

Changes in Nitrogen Deposition................................... 79

Regional Ecosystem Implications of Changes in Sulfur and Nitrogen 

Deposition ................................................. 81

Changes in the Ecological Condition of Lakes and Streams............... 80

Critical Loads ....................................................... 86

Summary . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .87

References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .89
v

 NAPAP Report to Congress: An Integrated Assessment

List of Tables

Table 1-1. Title IV NOx Affected Units by Boiler Type and NOx Emission Limit..... 7

Table 1-2. Estimated PM2.5 Health Benefits due to ARP Implementation in 2010 ... 15

Table 1-3. Estimated Ozone Health Benefits due to ARP Implementation in 2010 .. 15

Table 1-4. EGUs and SO2 Emissions Covered by Monitoring 

Method for the ARP, 2009 ........................................... 17

Table 2-1. Regional Changes in Air Quality from the 1989 to 1991 and 

2007 to 2009 Observation Periods ..................................... 25

Table 2-2. Regional Changes in Atmospheric Deposition from the 1989 to 1991 

and 2007 to 2009 Observation Periods .................................. 30

Table 2-3. Regional Trends in Sulfate, Nitrate, and ANC at 

LTM Sites from 1990 to 2008......................................... 35

Table 3-1. Comparison of Chronically and Episodically 

Acidified Streams in a Watershed....................................... 48

List of Figures

Figure 1–1. Trends in electricity generation, fossil fuel energy use, prices, and 

emissions from power plants affected by the ARP, 1990–2009 ................. 4

Figure 1-2. SO2 emissions and the allowance bank, 1995–2009 ................. 8

Figure 1-3. Cumulative SO2 allowances transferred under the ARP, 1994–2009 .... 9

Figure 1-4. SO2 allowances transferred under the ARP . ...................... 10

Figure 1-5. Average monthly SO2 allowance price, August 1994–May 2009 . ...... 11

Figure 1-6. Air quality, deposition, and water quality monitoring networks ....... 19

Figure 2-1. SO2 emissions from ARP sources, 1980 to 2009. .................. 22

Figure 2-2. Changes in annual SO2 emission levels by state from 1990 to 2009 

for ARP sources ................................................... 22

Figure 2-3. NOx emission levels for all ARP sources, 1990 to 2009. ............. 24

Figure 2-4. Changes in annual NOx emission levels by state from 1990 to 2009 

for ARP sources. ................................................... 24

Figure 2-5. National SO2 air quality, 1980 to 2009. ......................... 26

Figure 2-6. Annual mean ambient SO2 concentrations in the eastern United States 

for the 1989–1991 and 2007–2009 observation periods .................... 27

Figure 2-7. Annual mean ambient sulfate concentrations in the eastern 

United States for the 1989–1991 and 2007–2009 observation periods .......... 27

Figure 2-8. Annual mean ambient total nitrate concentrations in the eastern 

United States for the 1989–1991 and 2007–2009 observation periods .......... 29

Figure 2-9. Annual mean wet sulfate deposition in the United States for the 

1989–1991 and 2007–2009 observation periods. .......................... 31

Figure 2-10. Annual mean wet inorganic nitrogen deposition in the 

United States for the 1989–1991 and 2007–2009 observation periods. . ........ 33

Figure 2-11. Stages of recovery for acidic lakes and streams ................... 33
vi

Figure 2-12. LTM site locations and regional groupings ...................... 35

Figure 2-13. Trends in acidification of LTM lakes and streams, 1990–2008 ...... 36

Figure 2-14. Northeastern TIME Lakes by ANC Category, 1991–1994 vs. 

2006–2008 ....................................................... 38

Figure 2-15. Critical load exceedances in Adirondack Mountain 

TIME lakes, 1989 to 1991 (No. lakes = 117) ............................. 40

Figure 2-16. Critical load exceedances in Adirondack Mountain 

TIME lakes, 2006 to 2008 (No. lakes = 117) ............................. 40

Figure 2-17. Critical load exceedances in central Appalachian streams, 

1989 to 1991 (No. streams = 184) ..................................... 41

Figure 2-18. Critical load exceedances in central Appalachian streams, 

2006 to 2008 (No. streams = 184) ..................................... 42

Figure 2-19. 1999 to 2008 trends for the change in visibility on the clearest days 

(best 20% of visibility) as measured at 140 IMPROVE monitoring stations ...... 43

Figure 2-20. 1999 to 2008 trends for the change in visibility on the haziest days 

(worst 20% of visibility) as measured at 140 IMPROVE monitoring stations .... 44

Figure 3-1. Conceptual figure showing hypothetical acidification and recovery 

patterns in a waterbody that has been acidified by atmospheric deposition ....... 51

Figure 3-2. Map of major estuaries in the United States ...................... 56 

Figure 3-3. Climate change, acid deposition, and other stressors ................ 68

Figure 4-1. Stages of recovery for acidic lakes and streams ..................... 72

Figure 4-2. Map of acid-sensitive ecosystems in the United States ............... 74

Figure 4-3. Projected annual SO2 and NOx emission levels in 2020 

for the Base Case scenario and Scenarios A through C. Emission levels 

are for the continental United States ................................... 75

Figure 4-4. Projected change in annual total sulfur deposition from 

Base Case in 2020 with additional SO2 and NOx emission reductions .......... 77

Figure 4-5. Projected percentage change in annual total sulfur deposition from 

Base Case in 2020 with additional SO2 and NOx emission reductions .......... 78

Figure 4-6. Projected annual total nitrogen deposition from 

Base Case in 2020 with additional SO2 and NOx emission reductions .......... 79

Figure 4-7. Projected percentage changes in annual total nitrogen deposition 

from Base Case in 2020 with additional SO2 and NOx emission reductions ...... 80

Figure 4-8. Regions and waterbodies modeled using MAGIC .................. 83

Figure 4-9. Projected changes in the number of lakes and streams of 

acute concern (ANC < 0 μeq/L) in 2050 with additional SO2 and 

NOx emission reductions ............................................ 84

Figure 4-10. Projected changes in the number of lakes and streams of 

elevated concern (ANC 0–50 μeq/L) in 2050 with additional SO2 and 

NOx emission reductions ............................................ 85

Figure 4-11. Critical load exceedances in the Adirondack Mountains ............ 86
vii

 NAPAP Report to Congress: An Integrated Assessment

Acronyms and Abbreviations

AEL alternative emission limit 

AHM Alpine Hydrochemical Model

Al2+, Al3+ aluminum cations

AMS Allowance Management System 

AIRMoN Atmospheric Integrated Research Monitoring Network

ANC acid neutralizing capacity 

ARP Acid Rain Program 

BenMAP Benefits Mapping and Analysis Program

BTU British thermal unit 

CAAA Clean Air Act Amendments

CAIR Clean Air Interstate Rule 

CAMD Clean Air Markets Division

CASTNET Clean Air Status and Trends Network 

CAVR Clean Air Visibility Rule

CEMS continuous emission monitoring system

CENRS Committee on Environment, Natural Resources, 

and Sustainability

CFR Code of Federal Regulations

CMAQ Community Multi-Scale Air Quality

CMAS Community Modeling and Analysis System

CO2 carbon dioxide 

CPI consumer price index

CSAPR Cross-State Air Pollution Rule
viii

DOC dissolved organic carbon

DOE U.S. Department of Energy 

DOI U.S. Department of the Interior

dv deciview

ECMPS Emissions Collection and Monitoring Plan System 

EGUs electric generating units 

EMAP Environmental Monitoring and Assessment Program

EPA U.S. Environmental Protection Agency 

GAO U.S. General Accounting Office 

HNO3 nitric acid

ILWAS Integrated Lake-Watershed Acidification Study

IMPROVE Interagency Monitoring of Protected Visual Environments 

IPM Integrated Planning Model

kg/ha/yr kilograms per hectare per year

lb pound

LTM Long-Term Monitoring

MACT Maximum Achievable Control Technology

MAGIC Model of Acidification of Groundwater in Catchments 

meq/m2/yr milliequivalent per square meter per year

mmBTU million British thermal units

MPI market price index 

MW megawatt

N2O nitrous oxide

NAAQS National Ambient Air Quality Standards

NADP National Atmospheric Deposition Program

NAMS National Air Monitoring Stations 

NAPAP National Acid Precipitation Assessment Program
ix

 NAPAP Report to Congress: An Integrated Assessment

NASA National Aeronautics and Space Administration

NBP NOx Budget Trading Program

NDDN National Dry Deposition Network

NH3 ammonia

NH3 –N ammonia nitrogen

NH +

4 ammonium ion

NO nitric oxide

NO -

2 nitrite

NO -

3 nitrate

NO -

3 plus HNO3 ambient total nitrate

NO -

3 –N nitrate nitrogen

NOAA National Oceanic and Atmospheric Administration

NOx nitrogen oxides 

NOx SIP Call NOx State Implementation Plan Call

NPS National Park Service 

NSTC National Science and Technology Council

NSWS National Surface Water Survey 

NTN National Trends Network

O3 ozone

OMB Office of Management and Budget

OTC Ozone Transport Commission

PM2.5 fine particulate matter

PnET-BGC photosynthetic evapotranspiration-biogeochemical

RIA Regulatory Impact Assessment 

RSM Response Surface Model

RTC Report to Congress

SIP State Implementation Plan 
x

SLAMS State and Local Air Monitoring Stations 

SMBE simple mass balance equation 

SO2 sulfur dioxide 

SO 2-

4 sulfate

SSWC steady-state water chemistry

TIME Temporally Integrated Monitoring of Ecosystems

USDA U.S. Department of Agriculture

USGS U.S. Geological Survey

XML Extensible Markup Language 

μeq/L microequivalents per liter

μg/m3 micrograms per cubic meter 
EXECUTIVE SUMMARY

The National Acid Precipitation Assessment 

Program Report to Congress 2011

Acid deposition, more commonly known as acid 

rain, occurs when emissions of sulfur dioxide 

(SO2) and nitrogen oxides (NOx) react in the 

atmosphere with water, oxygen, and oxidants 

to form various acidic compounds. Prevailing 

winds transport the acidic compounds 

hundreds of miles, often across state and 

national borders. These acidic compounds 

then fall to earth in either a wet form (rain, 

snow, and fog) or a dry form (gases, aerosols, 

and particles). At certain levels, the acidic 

compounds, including small particles such as 

sulfates and nitrates, can cause many negative 

human health and environmental effects. 

What Are the Effects of Acid Rain?

Ecosystems and human health are subject to many 

stresses, including acid rain. Scientific research 

has shown that SO2 and NOx air pollutants and 

the acid rain formed by these pollutants can

•	 Degrade	air	quality,

•	 Impair	visibility,

•	 Negatively	impact	human	health,

•	 Acidify	lakes	and	streams,

•	 Harm	sensitive	forests,

•	 Harm	sensitive	coastal	ecosystems,	and

•	 Accelerate	the	decay	of	building	materials,	

paints, and cultural artifacts, such as 

buildings, statues, and sculptures.

Why Is This Report Being 

Sent to Congress, and What 

Is the Role of the National 

Acid Precipitation Assessment 

Program in the Report?

In	1990,	Congress	enacted	Title	IV	as	part	of	

the	Clean	Air	Act	Amendments	(CAAA).	The	

Acid	Rain	Program	(ARP)	created	under	Title	

IV	requires	significant	decreases	in	the	emissions	

of SO2 and NOx from fossil fuel–burning 

power	plants	to	improve	air	quality	and	protect	

ecosystems	that	have	suffered	damage	from	acid	

deposition.	Under	Title	IX	of	the	1990	CAAA,	the	

National	Acid	Precipitation	Assessment	Program	

(NAPAP)	was	asked	to	periodically	assess	and	

report	to	Congress	on	(1)	implementation	of	the	

ARP,	(2)	the	most	recent	scientific	information	

related	to	acid	deposition	and	its	effects,	and	(3)	

additional decreases in acid deposition necessary 

to	prevent	adverse	ecological	effects.	This	2011	

NAPAP	Report	to	Congress	(RTC)	focuses	on	

emission reductions from power plants, summarizes 

changes	in	deposition	rates	and	environmental	
Executive Summary | The National Acid Precipitation Assessment Program Report to Congress 2011 

ES 2

impacts,	and	projects	the	ecological	effects	of	

additional reductions in SO2 and NOx emissions.

What Are the Results of 

Implementing Title IV of the 1990 

Clean Air Act Amendments?

Implementation	of	Title	IV	has	successfully	

and substantially reduced emissions of SO2 and 

NOx	from	power	generation	(i.e.,	the	sources	

covered	by	the	ARP),	including	the	following.

•	 In	2009,	SO2 emissions were 5.7 million 

tons,	64%	lower	than	1990	emissions	and	

67%	lower	than	1980	emissions,	a	level	

below	the	2010	Title	IV	statutory	cap	of	

8.95	million	tons	of	SO2 emissions. 

•	 In	2009,	NOx emissions were 2 million 

tons,	67%	lower	than	1995	emissions,	

substantially	exceeding	the	Title	IV	goal	of	

a 2-million-ton reduction in NOx emissions 

from	projected	2000	levels	without	the	

ARP,	as	required	by	the	1990	CAAA.

•	 In	addition,	SO2 emissions from all sources, 

including	those	sources	not	covered	by	the	

ARP,	have	decreased	by	59%	since	1990,	

and emissions of NOx from all sources 

have	decreased	by	40%	since	1990.

The	emission	reductions	achieved	under	the	ARP	

have	contributed	to	measurable	improvements	

in	air	quality;	decreases	in	acid	deposition;	the	

beginnings	of	recovery	of	acid-sensitive	lakes	

and	streams	in	some	areas;	and	improvements	

in	visibility,	as	exhibited	by	the	following.

•	 SO2 concentrations in the atmosphere, a 

precursor to fine particles and acid deposition, 

have	decreased	since	1990.	Average	annual	

SO2	concentrations	in	2007–2009	were	

57%	to	63%	lower	than	in	1989–1991	in	

the Midwest and eastern United States.

•	 Sulfate	concentrations	in	the	atmosphere,	a	major	

component of fine particles, especially in the 

eastern	United	States,	have	decreased	since	1990.	

Average	annual	sulfate	concentrations	in	2007–

2009	were	41%	to	49%	lower	than	in	1989–

1991	in	the	Midwest	and	eastern	United	States.

•	 Wet	sulfate	deposition,	a	major	component	of	

acid	rain,	has	decreased	since	1990.	Average	

annual sulfate deposition in the Northeast and 

Southeast	in	2007–2009	was	43%	lower	than	in	

1989–1991,	deposition	in	the	Mid-Atlantic	was	

42% lower, and the Midwest was 44% lower.

•	 Wet	inorganic	nitrogen	deposition	has	

decreased	regionally	from	historical	levels.	

However,	decreases	were	less	than	those	

of wet sulfate deposition because of the 

continuing large contribution from other 

sources of NOx,	such	as	on-road	vehicles	and	

non-road	vehicles.	Still,	average	annual	wet	

inorganic	nitrogen	deposition	in	2007–2009	

was	16%	to	27%	lower	than	in	1989–1991	

in the Midwest and eastern United States.

•	 Levels	of	acid	neutralizing	capacity	(ANC),	

an indicator of the ability of a waterbody 

to	neutralize	acid	deposition,	have	shown	

improvement	from	1990	to	2008	at	many	lake	

and stream long-term monitoring sites in the 

eastern United States, including New England 

and	the	Adirondack	Mountains.	Many	lakes	and	

streams	still	have	acidic	conditions	harmful	to	

their	biota	even	though	the	increases	in	ANC	

indicate	that	some	recovery	from	acidification	

is	occurring	in	sensitive	aquatic	ecosystems. Water samples being collected on the Laurel Prong fork 

of the Rapidan River in Virginia (photo courtesy of U.S. 

EPA).
NAPAP Report to Congress: An Integrated Assessment

ES 3

•	 Data	from	the	Interagency	Monitoring	of	

Protected	Visual	Environments	(IMPROVE)	

network	show	no	statistically	significant	

trends	in	visibility	at	most	monitoring	sites	

over	the	10-year	period	from	1999	to	2008.	

The	Northeast,	however,	has	sites	with	

improving	visibility	on	both	the	best	and	worst

visibility	days,	principally	due	to	regionally	

decreased sulfate particulate concentrations.

Further, because emission reductions result in 

fewer fine particles and lower ozone concentration

in the air, thousands fewer premature deaths, 

hospital	admissions,	and	emergency	room	visits	

are	projected	in	2010.	The	value	of	the	resulting	

human health benefits from emission reductions 

range	from	$170	to	$430	billion	per	year	(2008$)

Emission	reductions	from	the	existing	ARP	have	

been augmented by additional programs, such 

as	the	Clean	Air	Interstate	Rule	(CAIR)	and	the	

Clean	Air	Visibility	Rule	(CAVR),	which	have	

increased	reductions	beyond	initial	ARP	levels.

What Is the Role of Cap and 

Trade in the Success of the 

Acid Rain Program?

The	success	of	the	SO2 emission-reduction progra

is	due	to	the	combined	use	of	an	overall	emission	

cap for SO2, which ensures that these reductions 

are	achieved	and	maintained,	and	a	trading	syste

that facilitates lowest-cost emission reductions. 

Together,	this	is	known	as	cap	and	trade.	NOx

reductions	under	the	ARP	are	achieved	through	a	

program that applies to a subset of coal-fired elect

generating	units	(EGUs)	and	is	closer	to	a	more	

traditional, rate-based regulatory system. Sources 

controlled in both the SO2 and NOx components

of	the	ARP	have	demonstrated	very	high	levels	of	

compliance,	averaging	99%	annual	compliance	

since	the	beginning	of	the	program.	The	inherent	

flexibility for sources to choose how to control the

SO2 emissions in the cap-and-trade approach for 

SO2 has been successful at reducing compliance 

costs	to	a	fraction	of	the	cost	estimated	in	1990.	

Several	factors	are	responsible	for	the	relatively	

low costs of SO2	reductions	realized	under	Title	

IV,	including	the	widespread	availability	of	low	

s 

.

m 

m

ric 

ir 

sulfur	coal	and	technical	innovations	that	facilitated	

use of that coal, lower than anticipated scrubber 

costs,	the	opportunity	to	bank	allowances,	and	

development	of	an	efficient,	high-volume	market	

for	allowances.	Although	the	costs	are	low,	the	

ARP	achieves	substantial	health	and	environmental	

benefits	through	air	quality	improvements.	

What Is the Future of 

Current Clean Air Rules?

Emissions of SO2 and NOx are expected to decline 

further as additional programs are implemented to 

control emissions from fossil fuel–burning power 

plants.	In	March	2005,	the	U.S.	Environmental	

Protection	Agency	(EPA)	promulgated	CAIR	to	

achieve	further	emission	reductions	beyond	levels	

reached	under	the	ARP	and	other	programs,	such	

as the NOx	State	Implementation	Plan	(SIP)	Call.	

CAIR	was	designed	to	help	states	in	the	eastern	

United	States	attain	the	National	Ambient	Air	

Quality	Standards	(NAAQS)	for	fine	particulate	

matter	(PM2.5)	by	reducing	and	capping	SO2

and NOx	emissions	in	28	eastern	states	and	the	

District	of	Columbia.	On	July	11,	2008,	the	U.S.	

Court	of	Appeals	for	the	D.C.	Circuit	issued	a	

ruling	vacating	CAIR	in	its	entirety.	The	Court	

subsequently	remanded	CAIR	to	EPA,	leaving	

CAIR	in	place	until	EPA	issued	new	rules	to	

replace	CAIR.	On	July	6th,	2011,	EPA	replaced	

CAIR	with	the	final	Cross-State	Air	Pollution	

Rule	(CSAPR),	which	will	control	SO2 and NOx

emissions from fossil fuel-burning power plants 

in	28	states	in	the	eastern	half	of	the	country.	

EPA	is	also	working	on	a	set	of	additional	rules	

covering	SO2 and NOx emissions from the electric 

power sector to address transport of pollution 

contributing to ozone and fine particle problems, as 

well	as	Maximum	Achievable	Control	Technology	

(MACT)	for	the	power	sector	and	other	industries.		

These	rules	could	lower	stationary	source	emissions	

for these same pollutants in the next 5 years.

Are Ecosystems Recovering 

from the Effects of Acid Rain?

Despite	the	environmental	improvements	reported	

here,	research	over	the	past	few	years	indicates	that	

recovery	from	the	effects	of	acidification	is	not	likely	
Executive Summary | The National Acid Precipitation Assessment Program Report to Congress 2011 

ES 4

for	many	sensitive	areas	without	additional	decreases	

in acid deposition. Many published articles, as 

well as the modeling presented in this report, 

show that the SO2 and NOx emission reductions 

achieved	under	Title	IV	from	power	plants	are	no

recognized	as	insufficient	to	achieve	full	recovery	

or	to	prevent	further	acidification	in	some	regions

Additional	SO2 and NOx emission reductions 

from power plants and other source sectors are 

needed	to	improve	air	quality,	reduce	deposition,	

and	further	reduce	the	number	of	acidic	lakes	

and streams in many regions of the United States.

Some of these additional emissions reductions 

may	be	achieved	through	implementation	of	

existing or future regulations to address transport 

of	ozone	and	fine	particles,	including	the	CSAPR	

in the eastern United States, and other rules 

affecting	mobile	sources,	SIPs,	NAAQS	for	ozone	

and	PM2.5, as well as future rules to reduce air 

toxics and other pollutants from power plants.

What Is the Importance of 

Long-term Environmental 

Monitoring in Understanding 

the Effects of Acid Rain?

Emissions,	air	quality,	deposition,	and	

ecological monitoring are critical components 

of	implementing	environmental	programs,	such	

as	the	Title	IV	ARP.	These	monitoring	efforts	

allow	researchers	and	policymakers	to	assess	the	

effectiveness	of	Title	IV	and	other	air	quality	

programs. Emissions monitoring is conducted by 

affected	sources;	additional	types	of	monitoring	ar

conducted	by	a	wide	variety	of	Federal	and	state	

agencies,	universities,	and	other	organizations.	

The	agencies	of	NAPAP	continue	to	have	a	strong

commitment to the research and monitoring that 

makes	assessments	like	this	NAPAP	RTC	possible

What Acid Rain–related Topics 

Are Currently at the Forefront 

of Scientific Knowledge?

Previous	NAPAP	RTCs	have	reported	on	the	

state of the science and emerging issues related 

to	acid	deposition.	This	report	also	summarizes	

recent	science	published	since	the	2005	NAPAP	

RTC.	Topics	covered	in	this	science	summary	

relate to ecosystem responses to emission 

w

.

e

.

controls and acid deposition reductions 

that	are	receiving	increasing	attention	in	the	

scientific literature, including the following:

•	 Scientists	have	observed	delays	in	

ecosystem	recovery	in	the	eastern	United	

States, despite decreases in emissions and 

deposition	over	the	past	30	years;

•	 A	growing	body	of	literature	documents	impacts	

of	nitrogen	deposition	on	western	ecosystems;

•	 “Critical	loads”	are	increasingly	used	by	

scientists	as	a	tool	for	quantifying	the	

sulfur	and	nitrogen	deposition	levels	at	

which ecosystems are impacted in order 

to	better	inform	air	quality	policies;

•	 Recent	literature	identifies	linkages	

between a changing climate, an altered 

carbon cycle and ecosystem response to 

acid deposition as an important emerging 

area	of	scientific	investigation;	and

•	 Scientists	identify	multi-pollutant	interactions,	

including reactions among SO2, NOx, ozone, and 

mercury, as an under-studied area of ecosystem 

impacts,	which,	if	better	understood,	could	have	

implications for future air pollutant policy.

Sugar maple is a tree species sensitive to acid rain 

(photo courtesy of U.S. EPA).
INTRODUCTION

The National Acid Precipitation Assessment 

Program

The National Acid Precipitation Assessment 

Program (NAPAP), a cooperative Federal 

program, was first authorized in 1980 and reauthorized under Title IX of the 1990 Clean 

Air Act Amendments (CAAA) to coordinate 

acid rain research and monitoring and to 

periodically report to Congress. NAPAP is 

comprised of the U.S. Environmental Protection 

Agency (EPA), the U.S. Department of Energy 

(DOE), the U.S. Department of Agriculture 

(USDA), the U.S. Department of the Interior 

(DOI)/U.S. Geological Survey (USGS), the DOI/

National Park Service (NPS), the National 

Aeronautics and Space Administration (NASA), 

and the National Oceanic and Atmospheric 

Administration (NOAA). The Acid Rain 

Program (ARP), authorized under Title IV of 

the CAAA, regulates the emissions of sulfur 

dioxide (SO2) and nitrogen oxides (NOx) 

from electric generating units (EGUs) that use 

fossil fuel (e.g., coal, gas, oil). These emissions 

contribute to acid deposition and the formation 

of fine particulate matter (PM2.5) and ozone, 

which ultimately leads to a wide range of 

environmental impacts, including harm to 

human health and visibility impairment. The 

NAPAP Reports to Congress (RTCs) provide an 

assessment of the implementation of the ARP, 

including the impacts and benefits of the sulfur 

and nitrogen emission reductions achieved by 

the program. This NAPAP RTC is written to 

effectively and fully communicate the results of 

the assessment to decision makers. Congress 

has asked NAPAP to assess all available data 

and information to answer two questions.

1. What are the costs, benefits, and 

effectiveness of Title IV? This question 

addresses the costs and economic impacts 

of complying with the ARP, as well as the 

benefit analyses associated with various 

human health and welfare effects, including 

reduced visibility and effects on ecosystems.

2. What reductions in deposition rates 

are needed to prevent adverse 

ecological effects? This is a complex 

question addressing how much 

deposition can occur before harmful 

environmental effects take place. 

Impacts of SO2 and NOx 

Emissions

SO2 and NOx emissions are the primary precursors 

involved in the formation of acid rain, also known 

as acid deposition. The ARP was created under 

the 1990 CAAA to reduce the adverse effects of 

acid deposition through reductions in annual 

emissions of SO2 and NOx. These emissions may 
Introduction | The National Acid Precipitation Assessment Program

 2

be transported up to hundreds of miles away from 

their emitting source, and thus have the potential to

impact large areas and populations. They may cause

such impacts as diminished air quality, damage to 

human health, acidification of lakes and streams, 

harm to sensitive forests and coastal ecosystems, 

degradation of visibility, and the acceleration of 

the decay of building materials. Certain acidsensitive ecosystems and regions may be more 

susceptible to the impacts of acid deposition. 

Additionally, some groups of people—such as 

children, the elderly, and people with pre-existing 

conditions—may be more sensitive to the related 

health impacts caused by diminished air quality 

due to SO2 and NOx emissions. The ARP regulates 

SO2 and NOx emissions from EGUs, but does 

not regulate vehicular or agricultural sources, both 

of which emit nitrogen and, to a lesser extent, 

sulfur compounds. In some areas, emissions 

from these other sources can be significant. 

Emissions have declined since the initiation of the 

ARP, and these reductions have led to increased 

visibility in some locations, reduced incidences 

of health impacts, improved conditions in some 

acidified lakes and streams, and reduced nitrogen 

deposition to some sensitive ecosystems. Ecological 

recovery has been identified in some areas, but this 

remains a complex process because recovery often 

lags behind declines in emissions. Current research 

looks to address the issue of delayed ecological 

response, as well as how climatic changes and 

multi-pollutant interactions may affect recovery.

Structure of the Report 

This NAPAP RTC is directed to Congress, but 

it provides valuable economic and scientific 

information to all public officials who are 

responsible for determining or evaluating air 

quality policy. The goal of this report is to present 

highly technical information pertinent to current 

public policy issues in a format that can be 

understood by the nonscientific reader. Where 

more scientific or economic detail is desired, 

references are noted in the text and provided at 

the end of the report. Chapters 1, 2, and 3 address 

the first question posed to NAPAP by Congress, 

and Chapter 4 focuses on the second question.

•	 Chapter	1	of	this	report	presents	the	status	

of implementation of Title IV, including 

information on ARP design, compliance, 

costs, and allowance transactions. 

•	 Chapter	2	presents	an	analysis	of	the	

observed changes, both past and present, in 

emissions of acid rain precursors, air pollutant 

concentrations, deposition of acidic species, and 

the measured effects of acidifying deposition 

on surface water quality and visibility. 

•	 Chapter	3	covers	advances	in	the	state	of	the	

science since the last NAPAP RTC regarding 

atmospheric deposition and the impacts of acid 

deposition on aquatic and terrestrial ecosystems, 

including ecosystems in the Rocky Mountains, 

the far western United States, and coastal areas. 

Chapter 3 also reports on research on the level 

of pollutant deposition below which ecosystem 

damage does not occur (i.e., critical loads), the 

interactions between atmospheric deposition and 

climate change, and multi-pollutant interactions. 

•	 Chapter	4	addresses	the	question	posed	by	

Congress in Title IX of the 1990 CAAA 

regarding the ecological impacts of further 

emission reductions and the uncertainties in 

the estimated relationships between emission 

reductions and the resulting ecological 

effects. Several scenarios representing a range 

of additional emission reductions are used 

to investigate the effects on acid-sensitive 

ecosystems in the eastern United States.

Visibility is good for hikers on Rescue Creek Trail in 

Yellowstone National Park (photo courtesy of NPS).
CHAPTER	1

Acid	Rain	Program	Elements		

and	Implementation

Established	under	Title	IV	of	the	1990	CAAA,	

the	ARP	requires	major	emission	reductions	

of	SO2	and	NOx,	the	primary	precursors	of	

acid	rain,	from	the	electric	power	industry.	

Since	its	implementation	in	1995,	the	ARP	

has	achieved	significant	emission	reductions	

as	electricity	generation	has	increased.	This	

chapter	focuses	on	the	ARP	and	includes	

descriptions	of	the	program	and	its	sources,	

program	compliance,	the	allowance	market,		

ARP	benefits	and	costs,	and	the	tools	used	

to	assess	the	progress	of	the	program.

Overview of Emission Reductions

The implementation of Title IV has successfully 

reduced emissions of SO2 and NOx from EGUs. 

Under Title IV of the 1990 CAAA, Congress 

established a permanent cap on the total amount 

of SO2 that may be emitted by EGUs in the 

conterminous United States. This cap has been 

phased in, with the final 2010 SO2 cap set at 8.95 

million tons—a level of about one-half of the 

emissions from EGUs in 1980. In 2009, 3,572 

EGUs were subject to the SO2 provisions of Title 

IV. By 2009, the sources (i.e., EGUs) covered 

by the ARP had reduced their combined SO2

emissions by approximately 67% from 1980 levels 

and 64% from 1990 levels. NOx reductions under 

the ARP apply to a subset of coal-fired EGUs and 

are regulated in a manner that is closer to a more 

traditional rate-based regulatory system. In 2009, 

the 960 sources subject to ARP NOx regulations 

emitted 67% fewer emissions than in 1995.1

Heat input and electricity generation increased 

by approximately 30% over this same period, and 

the average retail price of electricity was about 

the same in 2009 as it was in 1990 (Figure 1-1). 

The emission reductions achieved under Title IV are 

discussed in more detail in Chapter 2 of this report.

At a Glance: ARP Results in 2009

•	 SO2 emissions:	5.7	million	tons

• SO2 compliance:	100%

• SO2 allowances:	Allowance	bank	

increased	by	almost	4	million	allowances	

from	2008	levels

• SO2 allowance prices:	:	In	2009,	

allowance	prices	fell	from	$187	per	ton	to	

$61	per	ton

• NOx emissions:	2.0	million	tons

• NOx compliance:	100%

1 As described in Chapter 2 of this report, other programs—such as the Ozone Transport Commission (OTC); the NO Budget Trading Program x

(NBP) under EPA’s NOx State Implementation Plan (SIP) Call; and other regional and state NOx emission-control programs—also contributed 

significantly to the NOx reductions achieved by ARP sources in 2009. 
Chapter 1 | Acid Rain Program Elements and Implementation

4

-80%

-40%

-60%

1990 1992 1994 1996 1998 2000 2002 2004 2006 2008

Year

 SO2 Electricity Generation* Heat Input Emissions NOx

 Emissions Average Retail Price of Electricity**

* Generation from fossil fuel–red plants.

** Constant year 2000 dollars adjusted for ination.

Figure 1–1. Trends in electricity generation, fossil fuel energy use, prices, and emissions from power plants affected by 

the ARP, 1990–2009 (U.S. EPA, 2010b).

-20%

0%

20%

40%

60%

Percent Change

Phase I: 1995–1999 Phase II: 2000–2009

Source: Energy Information Administration (electricity generation, retail price); U.S. EPA (heat input and emissions, representing all affected ARP units). 

What Is Cap and Trade?

Cap	and	trade	is	a	policy	tool	for	protecting	human	health	and	the	environment	by	controlling	large	

amounts	of	pollution	from	a	group	of	sources.	A	cap-and-trade	program	first	sets	a	cap,	or	maximum	limit,	

on	pollution	emissions.	The	cap	is	chosen	in	order	to	achieve	a	desired	environmental	effect.	Sources	

covered	by	the	program	then	receive	authorizations	to	emit	in	the	form	of	emission	allowances,	with	

the	total	amount	of	allowances	limited	by	the	cap.	Each	source	can	design	its	own	compliance	strategy	

to	meet	the	overall	reduction	requirement.	For	example,	under	the	ARP,	sources	can	install	pollution	

controls,	implement	efficiency	measures,	change	to	lower	sulfur	coal,	or	sell	or	purchase	allowances,	among

other	options.	Individual	control	requirements	are	not	specified	under	a	cap-and-trade	program,	but	each	

emission	source	must	surrender	allowances	equal	to	its	actual	emissions	in	order	to	comply.	To	guarantee	

that	the	overall	cap	is	achieved,	sources	must	completely	and	accurately	measure	and	report	all	emissions.

Acid Rain Program Design

SO2 Program

The SO2 emission-reduction program created 

under Title IV represents a substantial change 

from traditional command and control regulatory 

approaches that establish source-specific emission 

limitations. Instead, the program combines an 

overall emission cap for SO2, which ensures that 

emission reductions are achieved and maintained, 

with a trading system that facilitates lowest-cost 

emission reductions. The ARP features tradable 

SO2 emission allowances, where one allowance 

is a limited authorization to emit one ton of 

SO2. A fixed number of allowances is issued by 

the government, and these allowances may be 

bought, sold, or banked for future use by EGUs 

or other parties (e.g., utilities, brokers, or anyone 

else interested in holding allowances). Existing 

sources are allocated allowances each year. New 

sources do not receive allowances and instead 

must buy them; however, the required purchase 

of allowances has not been a barrier to market 

entry (i.e., new sources have been able to acquire 
 NAPAP Report to Congress: An Integrated Assessment

5

the allowances needed to compete effectively in 

the market). At the end of each year, all affected 

sources are obliged to surrender to EPA the number 

of allowances that correspond to their annual SO2

emissions (one allowance for each ton of SO2). 

Title IV establishes a two-phased tightening of 

the SO2 emission cap, adjusting the allocation 

of SO2 allowances to fossil fuel–fired EGUs 

to reach the permanent cap on the number of 

allowances of 8.95 million tons annually in 

2010. Once the bank of unused allowances is 

depleted, the limit on allowance allocations 

constrains emissions to the level of the cap. 

Phase I of the ARP (1995–1999) affected 263 

of the larger (>100 megawatt [MW]), higheremitting EGUs, which are located primarily in 

the central and eastern United States. Phase I 

SO2 allowance allocations were distributed to each 

source based on the following formula: the product 

of an emission rate of 2.5 pounds (lb) SO2/million 

British thermal units (mmBTU) of heat input 

and its average heat input for 1985–1987. Some 

Phase II sources chose to “opt-in” to Phase I and 

comply early, bringing the total number of units 

participating in Phase I to more than 400. 

Phase II began in 2000 and extended to all 

existing EGUs serving generators larger than 

25 MW and all new fossil fuel–fired generation 

units throughout the country. In 2009, the total 

number of units covered by the SO2 criteria was 

3,572 sources, which represented 96% of total 

electricity generation from fossil fuels. In Phase II, 

all Phase I and Phase II SO2 affected sources are 

allocated allowances equivalent to an amount no 

greater than the product of 1.2 lb SO2/mmBTU 

and their average heat input for 1985–1987.

Title IV requires that sources monitor emissions 

continuously and report their emissions quarterly. 

Failure to surrender sufficient allowances results 

in two significant automatic penalties. Any source 

that fails to hold enough allowances to match its 

SO2 emissions for the previous year must pay to 

EPA, by July 1, an automatic penalty of $2,000 

(inflation-adjusted to $3,517 for 2009) per ton 

of emissions in excess of allowances held. The 

source must also immediately surrender to EPA 

an amount of allowances, issued for the year 

the payment is due, equaling the tons of excess 

emissions. A source may sell or bank for future 

use any remaining SO2 allowances not needed for 

compliance during a year. Sources may use these 

banked allowances as needed to comply with the 

program in future years until the bank is depleted.

Title IV mandates that a limited number of 

allowances allocable to existing sources be withheld 

and auctioned, with revenues from the auction 

returned pro rata to existing sources. The annual 

SO2 auction provides an opportunity for sources 

to buy and sell allowances. The auctions help 

ensure that new sources have an opportunity to 

obtain allowances beyond those allocated initially 

to existing EGUs. Complete results of the annual 

SO2 allowance auction are available at http://

www.epa.gov/airmarkets/trading/auction.html.

In addition to the Title IV SO2 program, recent 

programs such as the Clean Air Interstate 

Rule (CAIR) and the Clean Air Visibility Rule 

(CAVR) also have achieved SO2 reductions. 

Additional rules covering SO2 emissions from 

the electric power sector, as well as Maximum 

Achievable Control Technology (MACT) for 

the power sector and other industries, are being 

developed by EPA and may lead to further 

SO2 reductions from stationary sources.

Rampart Reservoir in Pike National Forest and the 

surrounding forests are examples of western habitats 

that may be impacted by acid precipitation (photo 

courtesy of U.S. EPA).
Chapter 1 | Acid Rain Program Elements and Implementation

6

NOx Program

In contrast to the system established for SO2

emissions, the ARP does not establish tradable 

emission allowances for NOx emission reductions. 

Instead, sources control how much NOx is emitted 

from coal-fired boilers based on the use of costeffective control technologies for each unit of fuel 

consumed (lb NOx per mmBTU). There are two 

phases of the NOx component: Phase I began in 

1996 (delayed 1 year because of litigation), and 

Phase II began in 2000. During Phase I, which 

applied to specific coal-fired boilers statutorily 

affected by Phase I SO2 requirements, the NOx

emission rate was set at 0.50 lb NOx/mmBTU for 

dry-bottom, wall-fired units and 0.45 lb NOx/BTU 

for tangentially fired units. Beginning in 2000, 

Phase II plants were required to meet emission 

rates between 0.40 lb NOx/mmBTU and 0.86 lb 

NOx/mmBTU, depending on the type of boiler. 

In 2009, 960 units were subject to ARP NOx 

program requirements, which represented 63% 

of total electricity generation from fossil fuels.

Although the ARP does not include NOx emission 

trading, sources are provided a degree of flexibility 

through emission-averaging provisions, whereby a 

company can meet its NOx emission limitations by 

averaging the emission rates of two or more boilers 

(see text box). This enables sources to reduce their 

NOx emissions at lower cost by allowing them to 

over-control at EGUs where it is technically easier 

to control emissions. At the end of the year, sources 

must demonstrate compliance with NOx emission 

requirements by achieving an annual emission 

rate at or below mandated levels, as outlined in 

their EPA-approved compliance plans. As with 

the SO2 program, it is important to note that a 

number of other programs have contributed to 

NOx emission reductions from ARP sources (see 

Chapter 2 for a description of these programs). 

Current programs, such as CAIR and CAVR, 

have achieved annual NOx emission reductions 

beyond those achieved by the Title IV NOx

program. Further reductions in NOx emissions 

from stationary sources will also come from the 

Cross-State Air Pollution Rule (CSAPR), which 

replaces CAIR to address the transport of pollution 

from the electric power sector contributing to 

ozone and fine particle problems, as well as MACT 

for the power sector and other industries.

Sources Achieved 100% NOx Compliance in 2009 Using 

a Variety of NOx Compliance Plan Options

Standard Limitation—A	unit	with	a	standard	limit	meets	the	applicable	individual	NOx	limit	prescribed	

for	its	boiler	type	under	40	CFR	Parts	76.5,	76.6,	or	76.7	(297	units	used	this	option	in	2009).

Alternative Emission Limit (AEL)—A	utility	can	petition	for	a	less-stringent	AEL	if	it	properly	installs	

and	operates	the	NOx	emission-reduction	technology	prescribed	for	that	boiler,	but	is	still	unable	to	

meet	its	standard	limit.	EPA	determines	whether	an	AEL	is	warranted	based	on	analyses	of	emission	data	

and	information	about	the	NOx	control	equipment	(three	units	used	this	option	in	2009).

Emissions Averaging—Many	companies	meet	their	NOx	emission-reduction	requirements	by	choosing	

to	become	subject	to	a	group	NOx	limit,	rather	than	by	meeting	individual	NOx	limits	for	each	unit.	The	

group	limit	is	established	at	the	end	of	each	calendar	year.	The	group	rate	must	be	less	than	or	equal	to	

the	British	thermal	unit	(BTU)–weighted	group	rate	that	the	units	would	have	had	if	each	had	emitted	at	

their	standard	limit	rate	(660 units	used	this	option	in	2009).

Note:		Unit	counts	do	not	include	those	with	a	retired	unit	exemption.
 NAPAP Report to Congress: An Integrated Assessment

7

Title IV Affected Sources

SO2 Sources

Sources that are subject to the SO2 component 

include boilers or combustion turbines that 

burn fossil fuel, serve generators with a design 

capacity greater than 25 MW, and produce 

electricity for sale. Several types of units meeting 

these criteria are not affected by the ARP. These 

include simple combustion turbines that began 

to produce electricity for sale before November 

15, 1990; cogeneration units whose annual 

electricity sales remain below the threshold 

established by regulation; and specific qualifying 

facilities and independent power producers that 

are contractually bound to sell electricity at a 

price that was established before November 15, 

1990. Despite these exceptions, almost all noncogeneration units that have total design capacity 

greater than 25 MW and that produce electricity 

for sale now must participate in the ARP. 

NOx Sources

Some of the sources subject to the SO2

requirements of Title IV are also subject to the 

Title IV NOx requirements. All units where coal 

accounted for more than 50% of heat input for at 

least 1 year during the 1990 through 1995 time 

period and that are configured for a specific type 

of boiler (i.e., cell burner, cyclone, dry-bottom 

wall-fired, tangentially fired, vertically fired, or wetbottom) are affected by the Title IV NOx criteria. 

Table 1-1 lists the sources affected by Title IV NOx

emission components in 2009. For more details on 

ARP applicability criteria, see 40 Code of Federal 

Regulations (CFR) 72.6, which provides ARP 

applicability regulations established under Title IV.

Compliance

SO2 Compliance

The Title IV SO2 program has achieved a nearperfect compliance record since the program took 

effect in 1995. In 2009, as in each year since 2005, 

the program achieved 100% compliance, and all 

ARP facilities complied with the requirement to 

hold enough allowances to cover SO2 emissions. 

EPA allocated 9.5 million SO2 allowances under 

the ARP for 2009. Together with the 8.5 million 

unused allowances carried over (or banked) from 

prior years, 18.0 million allowances were available 

for use in 2009 (see Figure 1-2). ARP sources 

emitted approximately 5.7 million tons of SO2 in 

2009, less than the allowances allocated for the 

year, and far less than the total allowances available

As a result, the bank increased between 2008 

and 2009 by nearly 4 million allowances to 12.3 

million, a 45% increase. The bank includes the 

unused allowances from previous years, plus the 

unused allowances allocated in 2009 (i.e., all of th

allowances above the yellow line in Figure 1-2). In 

2010, the total number of Title IV SO2 allowances

allocated annually dropped to 8.95 million and 

will remain statutorily fixed at that annual level.

Table 1-1. Title IV NOx Affected Units by Boiler Type and NOx Emission Limit

Coal-Fired	Boiler	Type

Title	IV	Standard	NOx	Emission	

Limits	(lb/mmBTU) Number	of	Units

Phase I, Group 1—Tangentially Fired 0.45 132

Phase I, Group 1—Dry-Bottom, Wall-fired 0.50 107

Phase II, Group 1—Tangentially Fired 0.40 296

Phase II, Group 1—Dry-Bottom, Wall-fired 0.46 290

Cell Burners 0.68 37

Cyclones >155 MW 0.86 54

Wet-Bottom > 65 MW 0.84 20

Vertically Fired 0.80 24

Total	All	Units 960

Source: U.S. EPA, 2009c

. 

e 
Chapter 1 | Acid Rain Program Elements and Implementation

8

The existence of the allowance market has given 

some sources the incentive to reduce their SO2

emissions below the level of their allowance 

allocation in order to sell their allowances to other 

sources or to “bank” them for use in future years. 

Other sources have been able to postpone or reduc

expenditures for pollution control by purchasing 

allowances from sources that controlled emissions 

beyond their allowance allocation level. As shown i

Figure 1-2, the bank or store of unused allowances

grew throughout Phase I (1995–1999) as sources 

reduced emissions more than required. These early

reductions reduced the amount of fine particles 

and acid deposition in the early years of Title IV 

implementation, increasing the human health 

and ecological benefits of the program in those 

years. Beginning in 2000, with Phase II of the 

program, the set of sources covered by the progra

expanded and the tighter Phase II emission cap 

took effect. As Figure 1-2 shows, sources began 

to use previously banked allowances in addition 

to allocations from the current year, to comply 

with Title IV. As a result, emission levels for 

2000–2005 were greater than annual allocations as

sources used banked allowances for compliance. 

e 

n 

m 

NOx Compliance

Affected sources can comply by either meeting a 

unit-specific NOx emission rate or including two 

or more units in an emission rate averaging plan. 

As with the SO2 program, the NOx program has 

had a high rate of compliance. Since 2000, there 

have been 4 years in which a single unit was out 

of compliance. In 2009, all 960 units that were 

subject to ARP NOx criteria achieved compliance.

2009 SO2 Allowance Market

The number of allowances (authorizations to emit

SO2) allocated to each source in any given year is 

determined by the Clean Air Act. A recent review 

(Burtraw and Szambelan, 2009) of emission tradi

generally concluded that the SO2 allowance mark

has “been liquid and active, and according to 

most observers, [has] worked well in achieving the

emission caps at less cost than would have been 

achieved with traditional approaches to regulation

Figure 1-3 shows the cumulative volume of 

SO2 allowances transferred under the ARP. The 

cumulative volume exceeded 400 million allowan

in 2009, with the majority of these allowances 

transferred in private transactions. The figure 

differentiates between allowances transferred in 

private transactions and those annually allocated 

and transferred to source accounts by EPA. 

ng 

et 

.”

ces 

Allowances Allocated

Allowances Banked in Previous Years

Actual Emissions from Affected Sources

Total Allowances Available for Use

SO2 Allowances (millions)

and Emissions (million tons) 

25

20

15

10

5

Phase I: 1995–1999 Phase II: 2000–2009

8.7

11.7

13.4

15.0

16.6

21.6

19.9

18.8 18.2

17.1 16.4 15.7 15.8 16.2

Year

1995 1996 1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008

10.2 9.4 8.9

7.6

10.6 10.2 10.6 10.3 11.2

4.9 5.3 5.4 5.5 5.3

18.0

2009

5.7

Figure 1-2. SO2 emissions and the allowance bank, 1995–2009 (U.S. EPA, 2010b).
 NAPAP Report to Congress: An Integrated Assessment

9

Private transactions are indicative of both market 

interest and the use of SO2 allowances as a 

compliance strategy. Of the nearly 406 million 

allowances transferred since 1994, about 67% were 

traded in private transactions. In December 2001, 

parties began to use a system developed by EPA 

to allow online SO2 allowance transfers. By 

2008, account holders registered over 99% of all 

private allowance transfers through EPA’s online 

transfer system. Allowance transfers are posted and 

updated daily on http://www.epa.gov/airmarkets.

In 2009, 2,716 private SO2 allowance transfers 

involving approximately 15.1 million allowances 

of past, current, and future vintages were recorded 

in EPA’s Allowance Management System (AMS). 

About 4 million SO2 allowances (26%) were 

transferred in economically significant transactions 

(i.e., between economically unrelated parties). 

Transfers between economically unrelated parties 

are “arm’s length” transactions and are considered 

a better indicator of an active, functioning market 

than are transactions among the various facility and 

general accounts associated with a given company. 

In the majority of all private transfers, allowances 

were acquired by power companies. Figure 1-4 

shows that the annual volume of SO2 allowances 

transferred under the ARP, excluding allocations, 

retirements (i.e., used allowances surrendered), 

and other transfers by EPA, has fluctuated since 

official recording of transfers began in 1994.

Phase I: 1995–1999 Phase II: 2000–2009

450

400

350

Allowances (millions)

300

250

200

150

100

50

1994 1995 1996 1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008 2009

Year

Private Transactions EPA Transfers to Account Holders

Figure 1-3. Cumulative SO2 allowances transferred under the ARP, 1994–2009 (U.S. EPA, 2010b).

SO2 Allowance Market in 

Brief (close of 2009)

Total Value of the SO2 Allowance 

Market: $1.1 billion*

•	 Average	Nominal	Price:	$61	per	ton

•	 Total	Allowance	Volume	(allowable	emissions):	

18,017,192

2009 Private Transactions

•	 2,716	transactions	moving	15.1	million	

allowances

•	 26%	of	allowances	transferred	between	

economically	unrelated	parties

*	Total	value	of	allowance	market	is	a	snapshot	

based	on	the	average	nominal	price	as	of	

December	2009	($61/ton)	and	the	total	allowance	

volume	available	for	2009	compliance.

Source:	U.S.	EPA,	2010b
Chapter 1 | Acid Rain Program Elements and Implementation

10

Figure 1-4. SO2 allowances transferred under the ARP (U.S. EPA, 2010b).

5

10

15

20

25

Trades Between Distinct Entities (signicant transfers)

Allowances (millions)

Trades Between Related Entities

30

0.9 1.9

4.4

7.9

9.5

6.2

12.7 12.6 11.6

8.1 7.5

10.0 9.5 9.1

5.9

9.2

16.7

8.2

15.2

13.5

18.7

25.0

22.5

21.4

16.5

15.3

19.9

22.4

16.9

13.9

4.0

15.1

1994 1995 1996 1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008 2009

Year

Phase I: 1995–1999 Phase II: 2000–2009

How Are Allowances Traded and Tracked?

Once	allowances	have	been	auctioned	and	allocated,	utilities	can	buy,	sell,	trade,	or	save	them	to	meet	their	

compliance	needs.	Along	with	the	utilities	that	hold	allowances	for	compliance	purposes,	other	parties,	such	

as	brokers,	environmental	groups,	and	private	citizens,	maintain	accounts	in	EPA’s	Allowance	Management	

System	(AMS).	The	AMS	database	records	account	balances	and	transaction	records	and	allows	public	access	

to	the	trading	history	of	each	allowance	until	it	is	finally	retired.	EPA	does	not	maintain	any	sensitive	business	

data,	such	as	the	price	associated	with	allowance	transfers.	Allowance	brokers	and	other	market	participants	

generally	maintain	a	market	price	index	(MPI)	to	track	trends	in	prices	over	time	and	provide	market	signals,	

similar	to	other	commodity	markets.

Most	allowance	transactions	take	place	in	the	over-the-counter	market,	where	prices	are	determined	by	

each	day’s	bids	and	offers	and	immediate	settlement	cash	trades	are	enacted	bilaterally	or	through	brokers.	

Once	trading	parties	agree	on	a	price,	they	generally	complete	the	transaction	using	standard	contracts	

developed	by	trade	associations	or	other	market	players	(see,	for	example,	the	sample	contract	available	at	

http://www.environmentalmarkets.org).	EPA	provides	a	list	of	brokers	and	environmental	groups	that	may	

be	interested	in	facilitating	trades	or	helping	parties	retire	allowances	voluntarily	(see	http://www.epa.gov/

airmarkets/trading/buying.html).	At	some	point	after	a	transaction	is	complete,	the	account	representative	of	

the	transferring	or	selling	party	will	usually	register	the	transfer	of	allowances	with	EPA.	The	representative	

can	submit	a	paper	form	or	transfer	the	allowances	online	using	the	Clean	Air	Markets	Division	(CAMD)	

Business	System	(see	http://www.epa.gov/airmarkets/business/transfer.	html).
 NAPAP Report to Congress: An Integrated Assessment

11

Over the first decade of the ARP, SO2 allowance 

prices were stable and significantly lower than 

projected. Just prior to the beginning of the 

program in 1995, SO2 allowance prices on the 

spot market were close to $150 per ton. The 

cost of allowances was initially projected to be 

between $250 and $500 per ton during Phase I 

(1995 to 1999) and $500 to $1,000 per ton in 

Phase II (beyond 2000). Actual allowance prices 

in Phase I were in the $100 to $200 range, with a 

low of $65 in 1996. Allowance prices did display 

some variability (as would commodities in any 

market), but it was within this very limited range 

and tended to be concentrated around times of 

regulatory change or uncertainty, such as the 

beginning of Phase I in 1995 and the transition 

to Phase II in 2000. Even as the more stringent 

Phase II requirements became effective in 2000, 

however, prices remained generally below the $200 

mark until they started to rise at the end of 2003 

with the proposal of CAIR (see Figure 1-5).

When CAIR was proposed in late 2003, allowance 

prices were influenced by the more stringent CAIR 

SO2 cap and new compliance deadline. After 

CAIR was finalized in March 2005, SO2 allowance 

prices continued to trend upward. CAIR was the 

most significant driver of the price adjustment that 

began in 2004 and culminated with prices around 

$1,600 per ton for a short time in December 

2005. The ARP SO2 market essentially became 

the CAIR SO2 market. The July 2008 court 

decision to vacate CAIR (subsequently revised to 

a remand in December of the same year) caused 

a dramatic drop in allowance prices as it created 

an uncertainty regarding the ongoing use of those 

allowances for compliance with the CAIR SO2

trading program requirements. During 2009, the 

value of the SO2 allowance market experienced a 

67% price decline; the monthly average allowance 

price fell from $187 per ton in January to $61 per 

ton by December (based on the market price index 

[MPI]) (see Figure 1-5). That decline continued 

in 2010, with the SO2 allowance price falling to 

an average of $40 per ton by May 2010. Together 

with the price decline, the volume of significant 

transactions fell sharply in 2009. Looking forward, 

it is estimated that allowance prices will remain 

low. Programs such as CAIR have led to significant 

reductions in SO2 emissions levels beyond the Title 

IV cap level. In addition, recent court decisions 

prohibit use of Title IV allowances for compliance 

in current or future programs, such as CAIR or 

the CSAPR, which replaces CAIR starting in 

2012. Thus, the Title IV SO2 trading program 

is likely to continue to be characterized by low 

volume and prices in the future. The program 

has met its environmental goals, and EPA expects 

that new emerging programs are likely to further 

lower power sector SO2 and NOx emissions.

Source: CantorCO2e, 2009 

$600

Nominal Price ($/ton)

$1,600

$1,800

$1,400

$1,200

$800

$1,000

$400

$200

$0

19941995

199619971998199920002001200220032004200520062007200820092010

Year

Phase I: 1995–1999 Phase II: 2000–2009

Figure 1-5. Average monthly SO2 allowance price, August 1994–May 2009 (U.S. EPA, 2009c).
Chapter 1 | Acid Rain Program Elements and Implementation

12

Market Changes

Market	observers	should	not	confuse	

temporary	high	prices	in	the	market	response

to	major	regulatory	changes	(i.e.,	more	or	

less	regulation)—where	buyers	and	sellers	

are	searching	for	a	new	equilibrium	based	

on	available	information	they	have	from	

consultants	and	various	services—with	

price	volatility.	EPA	and	market	analysts	have	

identified	these	regulatory	forces—the	CAIR	

emission	caps	and	compliance	deadlines,	

followed	by	the	rule	changes	resulting	from	

the	July	2008	CAIR	court	decision—as	the	

primary	factors	affecting	current	market	

conditions	in	the	period	2004–2008	and	not	

inherent	volatility	in	cap-and-trade	programs	

due	to	shifts	in	other	variables	that	influence	

the	market.	For	further	analysis,	see	http://

www.epa.gov/airmarkets/resource/docs/

marketassessmnt.pdf.

The Clean Air Interstate Rule

In	March	2005,	EPA	promulated	the	Clean	Air	Interstate	Rule	(CAIR)	to	achieve	further	emission	

reductions	beyond	levels	reached	under	the	ARP	and	other	programs,	such	as	the	NOx	State	

Implementation	Plan	(SIP)	Call.	CAIR	was	designed	to	help	states	in	the	eastern	United	States	address	

ozone	nonattainment	and	attain	the	National	Ambient	Air	Quality	Standards	(NAAQS)	for	PM2.5	by	

reducing	and	capping	SO2	and	NOx	emissions	in	28	eastern	states	and	the	District	of	Columbia.	CAIR	

created	three	separate	compliance	programs:	an	annual	NOx	program,	an	ozone-season	NOx	program,	and	

an	annual	SO2	program.	Each	of	the	three	programs	uses	a	two-phased	approach,	with	declining	emission	

caps	in	each	phase.	The	first	phase	began	in	2009	for	the	NOx	annual	and	NOx	ozone-season	program,	

and	started	in	2010	for	the	SO2	annual	program.	The	rule	also	establishes	a	second	phase	for	all	three	

programs,	beginning	in	2015.	CAIR	gave	affected	states	SO2	and	NOx	emission	budgets	and	the	flexibility	

in	their	SIPs	to	reduce	emissions	using	a	strategy	that	best	suits	their	circumstances,	including,	as	one	

option,	an	EPA-administered,	regional	cap-and-trade	program.	On	July	11,	2008,	the	U.S.	Court	of	Appeals	

for	the	D.C.	Circuit	issued	a	ruling	vacating	CAIR	in	its	entirety.	The	Court	subsequently	remanded	CAIR	

to	EPA	on	December	23,	2008,	leaving	CAIR	in	place	until	EPA	issues	new	rules	to	replace	CAIR.	On	July	6,	

2011,	EPA	finalized	the	Cross-State	Air	Pollution	Rule	(CSAPR),	which	replaces	CAIR	and	requires	eastern,	

southern,	and	central	states	to	significantly	improve	air	quality	by	reducing	power	plant	emissions	that	

cross	state	lines	and	contribute	to	ground-level	ozone	and	fine	particle	pollution	in	other	states.

Program Benefits and Costs

Benefits

Emissions of SO2 and NOx result in a variety of 

air pollutants, including not only the strong acids 

that impact ecosystems through atmospheric 

deposition but also atmospheric concentrations of 

particulate matter and ground-level ozone. These 

multiple air pollutants have important impacts 

on human health and a wide range of ecological 

and environmental resources. Due to the multipollutant nature of these environmental impacts, 

the emission reductions achieved by the ARP result 

in many societal benefits, including the following:

•	 Health Benefits. These include avoided 

premature mortality and avoided 

morbidity associated with reduced human 

exposures to PM2.5 and ozone, which 

are secondary air pollutants that form as 

a result of SO2 and NOx emissions.

•	 Visibility Benefits. Reductions in air pollutants, 

particularly in PM2.5, improve visibility, which 

leads to physical and economic benefits in 

both recreational and residential settings.
 NAPAP Report to Congress: An Integrated Assessment

13

•	 Agricultural and Forest Productivity Benefits.

Ground-level ozone inhibits plant growth. As a 

result, reductions in ozone concentrations yield 

physical and economic benefits in the form of 

enhanced agricultural and forest productivity.

•	 Ecological Benefits. A wide range of 

ecological resources are susceptible to damage 

when exposed to ambient air pollution or 

deposition of pollutants to terrestrial or aquatic 

environments. For a small portion of these 

effects, it is possible to quantify and estimate the 

economic value of avoided pollutant exposure.

•	 Materials Damage Benefits. Some materials 

are susceptible to accelerated deterioration when 

exposed to air pollution; as a result, reduction in 

air pollution can extend the life of these materials, 

yielding physical and economic benefits.

A comprehensive assessment of the benefits of 

ARP implementation would evaluate the entire 

suite of human health and environmental benefits 

resulting from ARP emission reductions. However, 

because human health impacts are more readily 

quantified and valued, air pollution benefits analyses 

traditionally have focused on human health rather 

than on ecological health, aesthetic effects, or 

natural resource productivity. For example, the 

science and economics of human health benefits 

assessment and valuation is much better developed 

than the corresponding science in support of 

assessing and valuing the monetary benefits 

of emission reductions on ecological systems. 

Moreover, the monetized human health benefits 

of reducing air pollution generally significantly 

outweigh monetary benefits in other categories, 

such as improvements in visibility or ecosystem 

condition. For these reasons, this report focuses 

on human health improvements in assessing the 

monetary benefits of ARP implementation.

Still, it is important to recognize that benefits 

beyond human health improvements result from 

emission reductions such as those achieved by 

the ARP. For example, a recent report (U.S. EPA, 

2011) fulfilling CAAA Section 812 requirements 

quantified the overall benefits of implementing 

the 1990 CAAA and estimated that all CAAA 

programs taken together will result in $40 billion 

R

i

in benefits due to improvements in recreational 

and residential visibility in 2010. There also have 

been recent advances in efforts to quantify the 

ecological benefits of emission reductions. For 

example, Banzhaf et al. (2006) used a contingent 

valuation study to estimate the total economic 

value of reducing ecological impacts of acid 

deposition in New York’s Adirondack Park from 

air quality policies that reduced SO2 and NOx

emissions. The study estimated total statewide 

benefits ranging from $336 million to $749 million 

annually. That total increased to $1.1 billion 

when alternative assumptions were used regarding 

ecological change. Significant future analytical 

work and basic ecological and economic research 

is needed to build a sufficient base of knowledge 

and data to support an adequate assessment 

of ecological benefits. For the current analysis, 

this incomplete coverage of effects represents a 

significant source of uncertainty in assessing the 

benefits of ARP implementation, which make up 

a significant portion of total CAAA benefits.

eductions in emissions of SO2 and NOx result 

n improved visibility (photo courtesy of NPS).
Chapter 1 | Acid Rain Program Elements and Implementation

14

Since publication of the last NAPAP RTC (NSTC,

2005), there have been some efforts to quantify 

the benefits of implementing the ARP. In 2005, a 

peer-reviewed journal article assessed the human 

health and welfare benefits of ARP implementation

for the prospective year, 2010 (Chestnut and 

Mills, 2005). The benefits were estimated using 

the modeled emission reductions and ambient air 

quality expected to be achieved in 2010 under the 

ARP. The majority of the monetized benefits of 

ARP implementation reported in the study were 

from the prevention of health-related impacts 

(e.g., premature death) due to reductions in ambie

concentrations of PM2.5 and ground-level ozone. 

A concentration response function developed by 

Pope et al. (2002) was used to estimate incidences 

of adult premature mortality as a result of PM2.5

exposure. EPA currently uses the Pope estimate to 

represent the lower end of the range of potential 

human health benefits. The Chestnut and Mills 

(2005) study estimated the PM2.5 and ozone health

related benefits of the ARP to be $134 billion and 

$5.5 billion (2008$) annually, respectively; they als

estimated benefits from visibility improvements in 

national parks and wilderness areas in California, 

the Colorado plateau, and the southeastern United 

States at about $2.5 billion (2008$) annually. 

Since publication of the Chestnut and Mills 

results, the assumptions used to develop human 

health effects estimates have changed. For example,

EPA now also includes concentration–response 

functions derived from a study by Laden et al. 

(2006) and an expert elicitation to estimate the 

range of incidences of adult premature mortality 

as a result of PM2.5 exposure. The Laden et al. 

estimate is currently used to represent the upper 

end of the range of potential human health 

benefits. Additionally, many underlying modeling 

assumptions have been updated, including 

population forecasts and baseline incidence 

rates. A majority of these updated assumptions 

are discussed in detail in the recent PM2.5

Regulatory Impact Analysis (U.S. EPA, 2006b).

nt 

-

o 

The benefits analysis included here updates the 

estimates found in the Chestnut and Mills (2005) 

study. This analysis was performed using EPA’s 

Benefits Mapping and Analysis Program (BenMAP) 

(U.S. EPA, 2008a). BenMAP is a tool that estimates 

the impacts of a change in air pollution on human 

health. Specifically, for this analysis, BenMAP 

was used to estimate the human health benefits 

from a reduction in PM2.5 and ground-level ozone 

pollution due to implementation of ARP emission 

reductions. The analysis relied on modeled air 

quality data representing expected air quality in 

2010, both in the absence of the ARP and with ARP 

implementation. BenMAP used these geographically 

distributed estimates of 2010 air quality to calculate 

a reduction in PM2.5 and ground-level ozone 

concentrations attributable specifically to ARP 

implementation. BenMAP combined this pollution 

reduction with geographically specific population 

data and baseline incidence data and entered 

this information into epidemiological functions 

to estimate health benefits. The epidemiological 

functions used for this analysis were the most 

recent sets of functions used by EPA for health 

effects assessments and employed in the PM2.5

Regulatory Impact Analysis (U.S. EPA, 2006b).

The results of the revised assessment show an 

increase in the estimated value of PM2.5 and ozone 

health benefits expected from ARP implementation 

in 2010. Depending on which of two studies is 

used (Pope et al., 2002 or Laden et al., 2006) 

as the primary estimate of incidences of adult 

mortality avoided, the range of monetized PM2.5

benefit increases ranges from 25% to 204% more 

than was estimated by Chestnut and Mills (2005) 

(see Table 1-2). Using updated methods to assess 

ground-level ozone benefits results in total benefits 

ranging from 75% to 319% of those estimated 

by Chestnut and Mills (2005) (see Table 1-3). As 

mentioned above, these updated benefits do not 

include human welfare benefits due to improved 

visibility or changed ecological conditions, such 

as reduced acidification of lakes and streams.
 NAPAP Report to Congress: An Integrated Assessment

15

a

 Valuation includes a 3% discount rate for future incidences of premature mortality avoided.

Note: Totals may not reflect the sum of individual rows due to rounding. 

Source: U.S. EPA, 2009b

Table 1-2. Estimated PM2.5 Health Benefits due to ARP Implementation in 2010

Health	Effect Incidences	Avoided Monetized	Value	

(millions;	2008$)

Adult Mortalitya

 from PM2.5

Pope et al., 2002 20,000 $160,000

Laden et al., 2006 50,000 $400,000

Range of Expert Elicitation 

Infant Mortality from PM2.5

Woodruff et al., 2006

Morbidity from PM2.5

Acute Bronchitis

7,000 to 66,000

82

28,000

$58,000 to $520,000

$710

$2.2

Acute Myocardial Infarction 30,000 $3,500

Acute Respiratory Symptoms 12,000,000 $790

Asthma Exacerbation 280,000 $15

Chronic Bronchitis 12,000 $5,800

Emergency Room Visits; Respiratory 18,000 $7.2

Hospital Admissions; Cardiovascular 10,000 $300

Hospital Admissions; Respiratory 4,800 $72

Lower Respiratory Symptoms 290,000 $5.6

Upper Respiratory Symptoms 220,000 $6.8

Work Loss Days

Total Value

Pope et al., 2002

2,500,000 $640

$170,000

Laden et al., 2006 $410,000

Morbidity

Hospital Admissions, Respiratory (age 65 and up) 3,000 $75

Hospital Admissions, Respiratory (age 0–2) 2,500 $26

Emergency Room Visits, Respiratory 1,900 $0.74

School Loss Day 910,000 $87

Acute Respiratory Symptoms 2,600,000 $170

Total Value Range $4,100–$17,000

Note: Totals may not reflect the sum of individual rows due to rounding. 

Source: U.S. EPA, 2009b

Table 1-3. Estimated Ozone Health Benefits due to ARP Implementation in 2010

Health	Effect Incidences	Avoided Monetized	Value	

(millions;	2008$)

Mortality

Mortality, Non-Accidental (Ito et al., 2005) 1,900 $17,000

Mortality, Non-Accidental (Schwartz, 2005) 660 $5,700

Mortality, Non-Accidental (Bell et al., 2004) 430 $3,700

Mortality, All Cause (Levy et al., 2005) 2,000 $17,000

Mortality, All Cause (Bell et al., 2005) 1,400 $12,000

Mortality, Cardiopulmonary (Huang et al., 2005) 720 $6,200
Chapter 1 | Acid Rain Program Elements and Implementation

16

Costs

In addition to the environmental benefits described 

above, significant economic benefits also have 

resulted from using the cap-and-trade mechanism 

employed by Title IV. Cap-and-trade programs 

provide sources with flexibility in how they achieve 

their emission target. The cap establishes the 

emission level for emission sources; however, the 

sources are provided with the flexibility of choosing 

how to abate their emissions. Each source can 

choose to invest in abatement equipment or energy 

efficiency measures, switch to fuel sources with 

no or reduced emissions, or shut down or reduce 

output from higher emitting sources. Cap-and-trade 

programs also allow sources to trade allowances, 

providing an additional option for complying 

with the emission target. Sources that have high 

marginal abatement costs (i.e., the cost of reducing 

the next unit of emissions) can purchase additional 

allowances from sources that have low marginal 

abatement costs. In this way, both buyers and sellers 

of allowances can benefit. Sources with low costs 

can reduce their emissions below their allowance 

holdings and earn revenues from selling their excess 

allowances—a reward for better environmental 

performance. Sources with high costs can purchase 

additional allowances at a price that is lower than 

the cost to reduce a unit of pollution at their facility.

Factors Responsible for the Low 

Cost of SO2 Reductions

•	 Switching to Low-sulfur Fuel.	Low-sulfur	

coal	became	less	costly	to	transport	at	the	

time	when	demand	increased.

•	 Low Scrubber Costs. Costs	were	lower	

than	expected.

•	 Technological Innovation. Technological	

improvements	that	allowed	switching	coals	

emerged	quickly.

•	 Efficient Allowance Market. 	An	efficient,	

high-volume	market	emerged.

•	 Banking of Unused Allowances. The	

program	offers	the	flexibility	to	bank	

allowances	for	future	use.

While this flexible emission-reduction framework 

is designed to maximize efficiency and minimize 

costs, some inefficiencies do exist. A recent analysis 

of Title IV implementation concluded that the 

program did not fully achieve least-cost and some 

opportunities for additional costs savings were 

unrealized, at least during the first several years 

of the program (Burtraw and Szambelan, 2009). 

Still, multiple studies on Title IV implementation 

have found that the program has had lowerthan-expected costs, particularly for Phase II, as 

well as cost savings compared to conventional 

regulatory approaches (Burtraw et al., 2005). 

In its 2005 RTC (NSTC, 2005), NAPAP 

reported on various cost estimates of Title IV 

implementation, including how the estimates had 

changed over time. Early projections of annual 

Phase I compliance costs ranged from just under 

$678 million (ICF, 1989) to $1.5 billion (EPRI, 

1993); later studies estimated that Phase I costs 

ranged from $940 million (Carlson et al., 2000) 

to $814 million (Ellerman, 2003)—the cost of 

emission reductions was found to be approximately 

50% lower than similar abatement under a 

command and control program (all estimates 

in 2000$). While assessments of Phase I costs 

show the cost effectiveness of reductions, an 

assessment of Phase II cost estimates shows that 

costs were much lower than initially projected. 

One of the first projections of Phase II costs from 

the Edison Electric Institute shows an expected cost 

of $7.5 billion annually in the year 2010. Similarly, 

the first EPA projection (1990) for annual Phase 

II costs was approximately $6 billion annually. 

However, as the approach of Phase I neared, 

projections for Phase II costs declined significantly. 

In 1994, the U.S. General Accounting Office 

(GAO) estimated Phase II costs to be approximately 

$2.5 billion per year in 2010 (2000$) (U.S. GAO, 

1994). Following that assessment, estimates 

provided by Ellerman (2003) and by Carlson 

et al. (2000) were $1.3 to $1.5 billion per year 

(2000$) and $1.1 billion per year (2000$) by 2010, 

respectively. An Office of Management and Budget 

(OMB) analysis presented in the 2005 NAPAP 

report estimated costs of the SO2 component to 

be between $1.1 and $1.8 billion (2000$) (NSTC, 
 NAPAP Report to Congress: An Integrated Assessment

17

2005). In the most recent study evaluating the 

cost of implementing only Title IV, Chestnut and 

Mills (2005) estimated total annualized costs at a 

slightly higher level than studies reported in the 

2005 NAPAP RTC (NSTC, 2005). Chestnut 

and Mills (2005) estimated the total annual costs 

for reducing SO2 at approximately $2 billion 

(2000$) per year, with NOx emission reductions 

costing an additional $1 billion annually—

still a fraction of the initial cost estimates.

Additionally, a report (U.S. EPA, 2011) under 

Section 812 of the CAAA analyzing the costs 

and benefits of implementing CAAA programs 

(Titles I through IV) estimates that the direct 

compliance costs in the year 2000 of implementing 

all programs affecting the utility sector was $1.37 

billion (2006$), or $1.17 billion when deflated to 

year 2000 (using the consumer price index [CPI] 

inflation calculator). Given that a substantial 

portion of the utility sector emission reductions 

under CAAA programs between 1990 and 2000 

resulted from Title IV implementation, this is a 

reasonable qualitative estimate of implementing 

Title IV through 2000. This estimate corroborates 

the estimates discussed above and included in the 

2005 NAPAP RTC (NSTC, 2005). The same 

Section 812 report (U.S. EPA, 2011) estimates 

costs of CAAA implementation in 2010 and 

2020. Those estimates show that the cost of 

CAAA program implementation increases in 

each of those years due to the implementation of 

CAIR and other emission-reduction programs. 

The costs to the government of administering 

the Title IV SO2 component are also less than in 

conventional regulatory programs. For example, the 

Title IV performance-based approach eliminates the 

need to devise source-specific emission limits, review 

control technologies, and prepare and approve 

detailed compliance schedules and permits. Because 

the regulating authority does not need to approve 

each source’s compliance choices, the focus is on 

ensuring that each source has at least one allowance 

for each unit of pollution emitted, which entails 

less administrative resources and expenditure.

Program Assessment Tools

Emission Monitoring and Reporting 

The ARP requires regulated sources to measure, 

record, and report emissions using continuous 

emission monitoring systems (CEMS) or an 

approved alternative measurement method. The vast 

majority of emissions are monitored with CEMS; 

however, alternatives are used at some facilities to 

provide an efficient means of monitoring emissions 

from the large universe of EGUs with lower overall 

mass emissions. Table 1-4 shows the number of 

units with and without SO2 CEMS for various 

fuel types in 2009, as well as the amount of SO2

emissions monitored using CEMS. Although only 

31% of units use CEMS, 99% of all SO2 emissions 

from ARP sources are monitored in this fashion.

Table 1-4. EGUs and SO Emissions Covered by Monitoring Method for the ARP, 2009

EGU	Type

2

Type	of	Monitoring	

System

Number	of	EGUs	

Monitored

Percentage	of	

EGUs	Monitored

CEMS 14 0.4 0.03

G

Percentage	of	SO2

Emissions	Monitored

Coal-Fired CEMS 1,044 29.4 98.8

as-Fired

Non-CEMS 2,277 64.1 0.06

CEMS 43 1.2 0.16

Oil-Fired

Non-CEMS 158 4.5 0.8

CEMS 13 0.4 0.15

Other

Non-CEMS 1 <0.1 <0.01

Note: The table excludes affected units that did not operate in 2009. “Other fuel units” include units that in 2009 combusted 

primarily wood, waste, or other nonfossil fuels.

Source: U.S. EPA, 2010b
Chapter 1 | Acid Rain Program Elements and Implementation

18

CEMS and approved alternatives are a cornerstone 

of the ARP’s accountability and transparency. 

Since the program’s inception in 1995, affected 

sources have met stringent monitoring quality 

assurance and quality control requirements and 

have reported hourly emission data in quarterly 

electronic reports to EPA. Using automated 

software audits, EPA rigorously checks the 

completeness, quality, and integrity of these 

data. All emission data are available to the public 

on the Data and Maps Web site maintained by 

EPA’s Clean Air Markets Division (CAMD) at 

http://camddataandmaps.epa.gov/gdm. Another 

CAMD Web site (http://www.epa.gov/airmarkets) 

provides access to other data associated with 

emission trading programs, including reports, 

maps, charts, and file downloads that cover source 

information, emissions, allowances, program 

compliance, atmospheric deposition and air 

quality, and aquatic ecosystem response.

Air Quality, Deposition, and 

Ecological Monitoring

Air quality, deposition, and ecological monitoring 

are also important components of the overall 

implementation of Title IV. This section presents 

information about the monitoring networks that 

are used to assess the progress of the ARP; other 

monitoring networks (e.g., National Air Monitoring

Stations [NAMS]) exist, but are not discussed 

here. Several monitoring networks designed to 

measure changes in air quality and acid deposition 

as a result of emission reductions are currently 

in operation and used by the ARP. In addition, 

surface water monitoring networks in acid-sensitive 

areas of the eastern United States measure changes 

in lake and stream chemistry in response to 

changes in emissions and atmospheric deposition. 

Together, this information allows policymakers 

to accurately assess the impact of Title IV and 

other air quality policies and to determine if the 

environmental goals are being achieved. Recent 

results of the ARP are presented in Chapter 2. 

Additionally, researchers continue to study the 

impacts of emission reductions on lakes, streams, 

forests, and coastal ecosystems (Chapter 3).

CASTNET and NADP/NTN monitoring site in 

Gunnison County, CO (photo courtesy of U.S. EPA).

Wet	Deposition	Monitoring

The National Atmospheric Deposition Program/

National Trends Network (NADP/NTN) is 

a nationwide network of predominantly rural 

precipitation monitoring stations. Operating 

since 1978, the network collects weekly wet acid 

deposition data to determine geographic patterns 

and temporal long-term trends. NADP/NTN 

is responsible for measuring the wet deposition 

component of total pollution loads across the 

United States. NADP/NTN is a collaborative 

effort between many different organizations and 

consists of 250 monitoring stations spanning 

the continental United States, Alaska, Canada, 

Hawaii, Puerto Rico, and the Virgin Islands 

(Figure 1-6). Additionally, the Atmospheric 

Integrated Research Monitoring Network 

(AIRMoN) is a NADP wet deposition subnetwork 

designed to quantify the extent to which changes 

in emissions affect air quality and atmospheric 

deposition at selected locations. AIRMoN-wet 

focuses on daily measurements of acids, nutrients, 

and base cations, including ammonium, in 

precipitation at seven sites. Quality assured data 

for all NADP networks are available from the 

NADP Web site at http://nadp.sws.uiuc.edu/.
 NAPAP Report to Congress: An Integrated Assessment

19

Air	Quality	and	Dry	Deposition	Monitoring

The Clean Air Status and Trends Network 

(CASTNET) is a regional, long-term environmental 

monitoring program administered and operated 

by EPA and NPS. Developed from the existing 

National Dry Deposition Network (NDDN), 

CASTNET was established in 1991 under the 

CAAA. The regional monitoring network was 

formed to assess trends in acidic deposition due 

to emission-reduction programs, such as the ARP. 

CASTNET has since become the nation’s primary 

monitoring network for measuring concentrations 

of air pollutants that form the dry component of 

acidic deposition and affect regional ecosystems 

and rural ambient ozone levels. CASTNET is able 

to provide the data needed to assess and report on 

geographic patterns and long-term temporal trends 

in ambient air pollution and dry atmospheric 

deposition. Presently, a total of 86 operational 

CASTNET sites are located in or near rural areas 

and sensitive ecosystems and collect data on 

ambient levels of pollutants where urban influences 

are minimal (Figure 1-6). As part of an interagency 

agreement, NPS sponsors 27 of the CASTNET 

sites, which are located in national parks and 

other Class-I areas designated as deserving special 

protection from air pollution. Quality assured data 

are available at http://www.epa.gov/CASTNET/.

Monitoring

Networks

CASTNET

IMPROVE

LTM-TIME

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Figure 1-6. Air quality, deposition, and water quality monitoring networks (prepared by U.S. EPA).
Chapter 1 | Acid Rain Program Elements and Implementation

20

Visibility	Monitoring

Interagency Monitoring of Protected Visual 

Environments (IMPROVE) is a long-term 

monitoring network that measures current visibility

conditions, tracks changes in visibility, and 

determines the causes of visibility impairment in 

national parks and wilderness areas. IMPROVE 

was established in 1985 to aid the development 

of Federal and state implementation plans to 

protect visibility in Class I areas, as stipulated in 

the 1977 CAAA. IMPROVE began collecting 

data in 1988 at 20 Class I areas. The network 

expanded to monitor the impacts of the 

Regional Haze Rule and now consists of 167 

sites nationwide (Figure 1-6). IMPROVE is 

a collaborative monitoring effort. Data and 

additional information about IMPROVE are 

available at http://vista.cira.colostate.edu/improve.

Visibility varies at the Look Rock Air Quality Monitoring 

Station in the Great Smoky Mountains National Park. 

On a good visibility day (top photo), one can see for 

100 miles. On a bad visibility day (bottom photo), the 

visual range shrinks to 15 miles (photo courtesy of 

NPS).

Ecological	Monitoring	–	Lakes	and	Streams

Temporally Integrated Monitoring of Ecosystems 

(TIME) and the Long-Term Monitoring (LTM) 

programs are complementary monitoring networks 

that provide information on a variety of indicators. 

These indicators are necessary for tracking temporal 

and spatial trends in environmental response to 

changes in regional air quality and acid deposition 

in ecosystems sensitive to acid rain in the eastern 

United States. TIME was developed as a special 

study within EPA’s Environmental Monitoring 

and Assessment Program (EMAP) to track trends 

in acid-relevant chemistry of particular classes of 

acid-sensitive lakes in the northeastern United 

States and streams in the central Appalachians. 

Because TIME sites were selected through a rigorous 

statistical sampling effort, measurements from 

these sites are used to extrapolate from a small 

number of regionally representative sampling sites 

to a much larger number of lakes and streams. 

In contrast, the primary objective of LTM is to 

detect long-term trends in the acid–base status of 

sensitive lakes and streams across a gradient of acidic 

deposition. The LTM network consists of a subset 

of lakes and streams that are particularly sensitive 

to acidity, with most site records extending back 

to the early 1980s. TIME and LTM monitoring 

sites are located in New England, the Adirondack 

Mountains, the Northern Appalachian Plateau, 

and the central Appalachians. Data are used to 

characterize how the most sensitive of aquatic 

systems in each region are responding to changing 

deposition and to provide information on 

seasonal chemistry and episodic acidification.
CHAPTER 2

Results of the Acid Rain Program: Status and 

Trends of Emissions, Deposition, Air Quality, 

Surface Water, and Visibility, 1990 to 2009

Since its inception in 1995, the ARP has made 

significant progress in reducing emissions. 

Starting in 2007, emissions were below the 

2010 SO2 cap set at 8.95 million tons, a level 

of about one-half of the emissions from EGUs 

in 1980. These emission reductions have led 

to important environmental benefits, including 

improvements in air quality; reductions in 

acid deposition; the beginnings of recovery 

from acidification in freshwater lakes and 

streams; and improvements in visibility.

Emissions

SO2 Emissions

As shown in Figure 2-1, SO2 emissions in the 

United States declined between 1980 and 2009 

(see data available at http://www.epa.gov/ttn/

chief/trends). In 2009, ARP sources had reduced 

annual SO2 emissions by 67%, compared with 

1980 levels, and 64%, compared with 1990 levels. 

ARP sources emitted 5.7 million tons of SO2

in 2009, which was well below the 2009 annual 

emission cap of 9.5 million tons and already below 

the annual cap of 8.95 million tons established 

by Title IV as the level for full implementation 

of the ARP in 2010. In addition, national SO2

emissions from all sources (including those 

not covered by the ARP) also have fallen by 

approximately 63%, from nearly 26 million tons 

in 1980 to about 9.5 million tons in 2009. The 

declines in SO2 emissions from all sources in the 

United States likely result from a combination 

of several factors, including the following: 

•	 Increased	use	of	emission-control	technologies,	

especially flue-gas desulfurization (or scrubbers);

•	 Reduced	heat	input	(a	measure	of	the	

amount of fuel used) at ARP sources; and

•	 Fuel	switching.

The SO2 requirements under the 

ARP apply to EGUs (i.e., fossil fuel –fired 

combustors) that serve a generator that 

provides electricity for sale. The vast majority 

of ARP SO2 emissions result from coal-fired 

EGUs, although the program also applies to 

oil and gas units.
The states with the highest-emitting sources 

in 1990 (i.e., Illinois, Indiana, Kentucky, 

Missouri, Ohio, Pennsylvania, Tennessee, and 

West Virginia) have generally seen the greatest 

SO2 reductions under the ARP (see Figure 2-2). 

Most of these states are upwind of the areas the 

ARP was designed to protect, and reductions 

have resulted in important environmental 

and health benefits over a large region.

Figure 2-1. SO2 emissions from ARP sources, 1980 to 2009 (U.S. EPA, 2010b).

Sources Covered All ARP Sources

by Phase II (2000 on)

Year

Phase I: 1995–1999 Phase II: 2000–2009

Phase I 

(1995–1999) Sources

Allowances Allocated

1980 1985 1990 1995 1996 1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008 2009

2

4

6

8

10

12

14

16

18

20

SO2 Emissions (million tons)

Total Emissions from 

All ARP Sources 

9.6 9.59.5 9.5 9.59.59.59.5 10.0

11.2

12.5 13.0 13.1 12.5 11.9

15.7

17.3

16.1

8.7 8.3

7.1 7.0 7.0

5.3

8.7

5.4 5.5 4.9 5.3

9.4 8.9

7.6

10.6 10.2 10.6 10.3 10.2

9.4 9.3

9.5

5.7

Figure 2-2. Changes in annual SO2 emission levels by state from 1990 to 2009 for ARP sources (U.S. EPA, 2010b).

1990 Emissions

1995 Emissions

2000 Emissions

2005 Emissions

2009 Emissions

Dashed line indicates that emissions remained 

below 400 tons per year from 1990–2009.

Scale: Largest bar equals 2.2 millions tons of 

SO2 emissions in Ohio, 1990.

Chapter 2 | Results of the Acid Rain Program

22
From 1990 to 2009, annual SO2 emissions in 41 

states and the District of Columbia fell by a total of 

approximately 10 million tons. In contrast, annual 

SO2 emissions increased by a total of 39,521 tons 

in seven states (Idaho, Montana, Nebraska, North 

Dakota, Oregon, Rhode Island, and Vermont) 

from 1990 to 2009. The seven states with the 

greatest reductions in annual emissions since 

1990 include Ohio, which decreased emissions by 

more than 1.5 million tons, and Illinois, Indiana, 

Kentucky, Pennsylvania, Tennessee, and West 

Virginia, each of which reduced total emissions 

during this time period by more than 500,000 

tons. To view emission data in an interactive 

format using Google Earth or a similar threedimensional platform, go to http://www.epa.gov/

airmarkets/progress/interactivemapping.html.

NOx Emissions

NOx emissions from all ARP sources were 2.0 

million tons in 2009, and emissions have decreased 

4.1 million tons since 1995 (see Figure 2-3). The 

goal of the ARP NOx program is to limit NOx

emissions from the affected coal-fired boilers so 

that their emissions are at least 2 million tons 

less than the projected level for the year 2000 

without implementation of Title IV. The 2009 

emission level of 2.0 million tons is 6.1 million 

tons less than the projected level of annual NOx 

emissions in 2000 without the ARP, or more 

than triple the Title IV NOx emission-reduction 

objective. While the ARP was responsible for a 

large portion of these annual NOx reductions, 

other programs—such as the NOx Budget 

Trading Program (NBP) under EPA’s NOx State 

Implementation Plan (SIP) Call and other regional 

and state NOx emission-control programs—also 

contributed significantly to NOx reductions.

As noted, from 1995 to 2009, annual NOx 

emissions from ARP sources dropped by 4.1 

million tons, a net decrease of 67%. During 

this period, 44 states and the District of 

Columbia reduced NOx emissions, while 6 other 

states accounted for only about 5,542 tons of 

increased NOx emissions (see Figure 2-4). 

Other Programs Contributing to NOx Emission Reductions

Ozone Transport Commission (OTC)—Established under the 1990 CAAA, this commission consists of 

states primarily located in the Northeast and mid-Atlantic region. The OTC developed the OTC NOx Budget 

Program, which operated from 1999–2002. As part of this program, 11 states and the District of Columbia 

entered into a memorandum of understanding to achieve regional emission reductions of NOx through the 

use of control technologies and an ozone-season cap-and-trade program.

NOx SIP Call—Issued in 1998 to reduce the regional transport of ground-level ozone, the NOx SIP Call 

required states to reduce ozone-season NOx emissions by meeting emission budgets.

NOx Budget Trading Program (NBP)—This market-based cap-and-trade program was developed under 

the NOx SIP Call and replaced the OTC NOx Budget Program in 2003. The NBP was created to reduce NOx

emissions from power plants and other large combustion sources in the eastern United States.

Clear Air Interstate Rule (CAIR)—Promulgated in 2005, CAIR was designed to reduce emissions of SO2

and NOx and replaced the NBP in 2009. This rule created three separate trading programs: an annual NOx

program, an ozone-season NOx program, and an annual SO2 program. In 2009, CAIR was remanded, but 

remained in place while EPA developed a new rule (U.S. EPA, 2009e).

Cross-State Air Pollution Rule (CSAPR)—On July 6, 2011, EPA finalized the CSAPR, which replaces 

CAIR and requires eastern, southern, and central states to significantly improve air quality by reducing power 

plant emissions that cross state lines and contribute to ground-level ozone and fine particle pollution in 

other states. The rule reduces annual SO2, annual NOx, and ozone-season NOx emissions, while providing 

sources with flexibility in how to comply with the program. The rule also allows air quality-assured trading 

that uses an allowance market infrastructure based on existing, successful allowance trading programs. At the 

same time, the trading ensures the elimination, in each state, of the emissions that contribute to downwind 

air quality problems.

23

NAPAP Report to Congress: An Integrated Assessment
Figure 2-4. Changes in annual NOx emission levels by state from 1990 to 2009 for ARP sources (U.S. EPA, 2010b).

Year

6.1

6.7

6.0 5.9

5.5

6.0

4.7

5.1

4.2

4.5

3.6 3.8

3.3 3.4

3.0

5.5 5.4 5.4 5.5

4.8

5.3

4.1

4.5

3.8 4.0

3.3 3.4

3.1 3.0 2.7

1990 1995 1996 1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008

Total NOx Emissions 

from all ARP Sources 

2.7

1

2

3

4

5

6

7

8

9

NOx Emissions (million tons)

NOx Emissions from ARP 

Sources Covered for SO2

but not for NOx

NOx Emissions from ARP 

Sources Covered for NOx

1990 Projection of NOx

Emissions from ARP Sources 

in 2000

Phase I: 1995–1999 Phase II: 2000–2009

1.8

2.0

2009

Figure 2-3. NOx emission levels for all ARP sources, 1990 to 2009 (U.S. EPA, 2010b).

Chapter 2 | Results of the Acid Rain Program

24

1990 Emissions

1995 Emissions

2000 Emissions

2005 Emissions

2009 Emissions

Scale: Largest bar equals 500,000 tons of 

NOx emissions in Ohio, 1990.

Note: Darker area represents the NOx

Budget Trading Program geographic region.

Dashed line indicates that emissions remained 

below 400 tons per year from 1990–2009.
The states subject to EPA’s 1998 NOx SIP Call 

have achieved significant reductions in ozoneseason NOx emissions since the baseline years of 

1990 and 2000. All of these states have achieved 

reductions since 1990 as a result of programs 

implemented under the 1990 CAAA, with many 

states reducing their emissions by more than half 

since 1990. A significant portion of these decreases 

in NOx emissions has been achieved since 2000, 

largely as a result of decreases under ozone-season 

NOx trading programs implemented by the Ozone 

Transport Commission (OTC) from 1999 to 2002,

and under the NOx SIP Call from 2003 to 2009. 

For reports about these programs, go to http://www

epa.gov/airmarkets/progress/progress-reports.html. 

Air Quality

Emission reductions achieved under the ARP 

have led to improvements in air quality, with 

significant benefits to human health. Since the early

1990s, improvements in ambient SO2, sulfate, 

and nitrate concentrations have varied regionally 

in the eastern United States (see Table 2-1).

.

Table 2-1. Regional Changes in Air Quality from the 1989 to 

1991 and 2007 to 2009 Observation Periods

Region Average Concentration 

(μg/m3), 1989–1991a

Average Concentration 

(μg/m3), 2007–2009a

Percentage 

Changeb

Number 

of Sites

Ambient SO2

Northeastc 5.5 1.7 -69 3

Mid-Atlantic 13.0 5.0 -62 12

Southeast 5.1 2.2 -57 9

Midwest 11.0 4.1 -63 10

Ambient Sulfate

Northeastc 3.5 1.8 -49 3

Mid-Atlantic 6.3 3.5 -44 12

Southeast 5.4 3.2 -41 8

Midwest 5.8 3.1 -47 10

Total Ambient Nitrate 

Northeastc 1.8 1.0 -44 3

Mid-Atlantic 3.3 2.0 -39 12

Southeast 2.2 1.5 -32 8

Midwest 4.6 3.2 -30 10

a Averages are the arithmetic mean of all sites in a region that were present and met the completeness criteria in both averaging 

periods. Thus, the average concentrations may differ from past reports.

b All values are statistically significant at the 95% confidence level unless otherwise indicated.

c Percentage change in this region was not tested for statistical significance because too few monitoring sites are available.

Source: U.S. EPA, 2010c

SO2 and Sulfate

Data collected from monitoring networks show 

that the decline in SO2 emissions from the power 

industry has improved air quality. Based on 

EPA’s latest air emission trends data (see http://

www.epa.gov/airtrends/index.html), the national 

composite average of SO2 annual mean ambient 

concentrations decreased 76% between 1980 

and 2009, as shown in Figure 2-5 (based on 

state, local, and EPA monitoring sites located 

primarily in urban areas). Although Figure 2-5 

shows a steady declining trend from 1980–2009, 

the largest single-year reduction (20%) occurred 

in the first year of the ARP, between 1994 and 

1995. The second-largest single-year reduction 

(16%) occurred most recently between 2008 and 

2009. These trends are consistent with the regional 

ambient air quality trends observed in CASTNET. 

During the late 1990s, dramatic regional 

improvements in SO2 and ambient sulfate 

concentrations were observed at CASTNET sites 

throughout the eastern United States following 

25

NAPAP Report to Congress: An Integrated Assessment
implementation of Phase I of the ARP. These 

improvements continue today. Analyses of regional 

monitoring data from CASTNET show the 

geographic pattern of SO2 and airborne sulfate 

in the eastern United States. Three-year mean 

annual concentrations of SO2 and sulfate from 

CASTNET long-term monitoring sites in the 

eastern United States are compared for the 1989 

to 1991 and 2007 to 2009 observation periods 

(see Figures 2-6 and 2-7). For the 1989 to 1991 

observation period, few data on the ambient 

concentrations of SO2 and sulfate exist from 

CASTNET sites in the western United States; 

therefore, changes in ambient concentrations in this 

region could not be assessed for these two periods.

Figure 2-6 shows that from 1989 to 1991, prior to 

implementation of Phase I of the ARP, the highest 

annual ambient concentrations of SO2 in the eastern 

United States were observed in western Pennsylvania 

and along the Ohio River Valley. In comparison, 

the map for the 2007 to 2009 observation period 

indicates a significant decline in ambient SO2

concentrations in nearly all affected areas after 

implementation of the ARP and other programs.

Figure 2-5. National SO2 air quality, 1980 to 2009 (U.S. EPA, 2010c). SO2 Ambient Concentration (ppm)

Average concentration 10% of sites have 

concentrations below this line 

90% of sites have concentrations

below this line

National Ambient Air

Quality Standard 

0

0.01

0.02

0.03

0.04

Phase I:

1995–1999

Phase II:

2000–2009

Year

1980 1982 1984 1986 1988 1990 1992 1994 1996 1998 2000 2002 2004 2006 2008 2009

Concentration—A measure of how much of 

a given substance (e.g., nitric acid) is mixed with 

another substance (e.g., water), expressed as a 

quantity or mass per unit volume.

Deposition—A measure of the amount of 

chemical that is transferred to the earth’s 

surface—typically by rainfall, snow, hail, or 

sleet—expressed as the amount or mass 

transferred to a specific area of ground.

Deposition is a combination of both concentration

and the amount of rain. For example, the West has

a higher atmospheric pollutant concentration, but 

less rainfall; the East has a lower concentration, but 

much more rain and therefore more deposition.

Note: Example is only for wet deposition (courtesy of NADP).

Chapter 2 | Results of the Acid Rain Program

26
Figure 2-6. Annual mean ambient SO2 concentrations in the eastern United States for the 1989–1991 and 

2007–2009 observation periods (U.S. EPA, 2010c).

0

2

4

6

8

10

12

14

16

18

> 20

SO2

(μg/m3)

0

2

4

6

8

10

12

14

16

18

> 20

SO2

(μg/m3)

Note: For maps depicting these trends for the entire conterminous United States, visit http://www.epa.gov/castnet. 

Dots on all maps represent monitoring sites. Lack of shading for southern Florida on the map on the left indicates lack 

of monitoring coverage in the 1989 to 1991 time period.

1989 to1991 2007 to 2009

27

NAPAP Report to Congress: An Integrated Assessment

0.0

1.0

2.0

3.0

4.0

5.0

6.0

7.0

> 8.0

SO4

2-

(μg/m3)

0.0

1.0

2.0

3.0

4.0

5.0

6.0

7.0

> 8.0

SO4

2-

(μg/m3)

1989 to1991 2007 to 2009

Note: For maps depicting CASTNET data for the conterminous United States, visit http://www.epa.gov/castnet. 

Dots on all maps represent monitoring sites. Lack of shading for southern Florida on the map on the left indicates lack 

of monitoring coverage in the 1989 to 1991 time period.

Figure 2-7. Annual mean ambient sulfate concentrations in the eastern United States for the 1989–1991 and 

2007–2009 observation periods (U.S. EPA. 2010c).
What is the Difference Between SO2, sulfate, and total sulfur deposition?

Sulfur is an element that exists in several different forms. SO2 and sulfate are the forms of sulfur examined 

in this report. 

SO2—When sulfur-containing substances, such as coal, are burned, the sulfur is primarily converted to 

SO2. Emissions from ARP sources and ambient air quality are discussed in terms of SO2 concentrations.

Sulfate—SO 2- 2 is oxidized in the atmosphere to form sulfate (SO4 ). Acid produced during the oxidation process is a major contributor to wet acidifying deposition. In this report, ambient air quality and 

levels of wet deposition are discussed in terms of sulfate concentrations. Also, sulfate concentrations in 

water are an indicator (along with base cation and nitrate concentrations) of lake and stream acidification where there are no easily weathered soil or bedrock sources of sulfate (e.g., gypsum).

Total Sulfur—Total sulfur represents the sum of all sulfur species. In this report, total deposition is 

discussed in terms of the levels of sulfur deposited, representing the sum of wet and dry deposition.

Like SO2 concentrations, ambient sulfate 

concentrations have decreased since the ARP 

was implemented, with average values decreasing 

by 41% to 49% throughout the eastern United 

States. During the 1989 to 1991 observation 

period, the highest annual ambient sulfate 

concentrations were observed in western 

Pennsylvania, along the Ohio River Valley, 

and in northern Alabama at levels greater than 

11 micrograms per cubic meter (μg/ m3). Most 

of the eastern United States experienced annual 

ambient sulfate concentrations greater than 

5μg/ m3. Since the ARP was implemented, both 

the magnitude and spatial extent of the highest 

ambient sulfate concentrations have dramatically 

declined, with the largest decreases observed 

along the Ohio River Valley (see Figure 2-7).

Atmospheric deposition occurs as wet 

deposition, which falls to the earth through 

rain, snow, and fog, and as dry deposition, 

which falls to the earth as gas and particles 

in the absence of precipitation or fog. Total 

atmospheric deposition is the sum of wet and 

dry deposition.

Chapter 2 | Results of the Acid Rain Program

28

NOx and Nitrate

Although the ARP has met its NOx emissionreduction targets, emissions from other 

sources (e.g., mobile sources) contribute to 

ambient nitrate concentrations in many areas. 

NOx levels also can be affected by emissions 

transported via air currents over wide regions.

From 2007 to 2009, reductions in NOx emissions 

during the ozone season from power plants under 

the NOx SIP Call and CAIR have continued to 

result in significant region-specific improvements 

in ambient total nitrate (i.e., nitrate [NO -

3 ] plus 

nitric acid [HNO3]) concentrations. For instance, 

annual mean ambient total nitrate concentrations 

for the 2007 to 2009 observation period in the 

mid-Atlantic region were 39% less than the annual 

mean concentration in the 1989 to 1991 period 

(see Figure 2-8). Although these improvements 

might be partly attributed to added NOx controls 

installed for compliance with the NOx SIP Call and 

CAIR, the findings at this time are not conclusive.
Figure 2-8. Annual mean ambient total nitrate concentrations in the eastern United States for the 1989–1991 and

2007–2009 observation periods (U.S. EPA, 2010c).

0.0

1.0

2.0

3.0

4.0

5.0

> 6.0

Total NO3

-

(μg/m3)

1989 to1991 2007 to 2009

Note: For maps depicting these trends for the entire conterminous United States, visit http://www.epa.gov/castnet. 

Dots on all maps represent monitoring sites. Lack of shading for southern Florida on the map on the left indicates lack 

of monitoring coverage in the 1989 to 1991 time period.

0.0

1.0

2.0

3.0

4.0

5.0

> 6.0

Total NO3

-

(μg/m3)

What is the difference between NOx, nitrate, and total nitrogen deposition?

Nitrogen exists in several different forms that are of interest in this report, including NOx and nitrate. 

NOx—During combustion, nitrogen in fuel (e.g., coal) and the atmosphere combines with oxygen at high 

temperatures and pressure to form NOx. Emissions from ARP sources are discussed in terms of NOx

concentrations.

Nitrate—NO is oxidized in the atmosphere to form nitrate (NO - x 3 ). Similar to SO2, acid produced during 

the oxidation process is a contributor to acidifying deposition. In this report, ambient air quality and wet 

deposition are discussed in terms of nitrate concentrations. Total ambient nitrate concentrations reported 

here consist of the sum of nitric acid, ionic nitrates, and particulate nitrates. Also, nitrate concentration in 

water is an indicator (along with base cation and sulfate concentrations) of lake and stream acidification.

Total Nitrogen—Total nitrogen represents the sum of all nitrogen species. In this report, total nitrogen 

deposition is discussed in terms of the levels of nitrogen deposited, representing the sum of wet and dry 

deposition.

29

NAPAP Report to Congress: An Integrated Assessment
Acid Deposition

NADP/NTN monitoring data show significant 

improvements across the eastern United States in 

the primary acid deposition indicators—sulfur 

and nitrogen. The sulfur indicator is assessed using 

the atmospheric deposition levels of sulfate (wet 

deposition) and sulfur (dry and total deposition). 

The nitrogen indicator is measured using levels 

of wet inorganic nitrogen deposition (combined 

deposition of inorganic nitrate and ammonium 

ions in wet deposition) and total inorganic 

nitrogen deposition (combined deposition of 

inorganic nitrate and ammonium deposition in 

precipitation, dry particulate deposition, and 

gaseous nitric acid). Table 2-2 provides an overview 

of changes in atmospheric deposition in regions 

of the eastern United States from the 1989 to 

1991 and 2007 to 2009 observation periods.

Table 2-2. Regional Changes in Atmospheric Deposition from the 

1989 to 1991and 2007 to 2009 Observation Periods

Region Average Deposition 

(kg/ha), 1989–1991

Average Deposition 

(kg/ha), 2007–2009

Percent

Changea

Number

of Sites

Wet Sulfate 

Northeast 7.5 4.3 -43 17

Mid-Atlantic 9.2 5.3 -42 11

Southeast 6.1 3.5 -43 23

Midwest 7.1 4.0 -44 27

Dry Sulfur

Northeastb 2.9 1.0 -66 3

Mid-Atlantic 6.7 2.9 -57 11

Southeastb 1.2 0.7 -42 2

Midwest 6.5 2.8 -57 10

Total Sulfurc

Northeastb 9.8 4.7 -52 3

Mid-Atlantic 16 8.0 -50 11

Southeastb 8.0 4.6 -43 2

Midwest 15.0 7.0 -53 10

Wet Inorganic Nitrogen

Northeast 5.6 4.1 -27 17

Mid-Atlantic 6.2 4.5 -27 11

Southeast 4.4 3.4 -23 23

Midwest 5.8 4.9 -16 27

Dry Inorganic Nitrogen

Northeastb 1.4 0.6 -57 3

Mid-Atlantic 2.5 1.5 -40 11

Southeastb 0.9 0.8 -11 2

Midwest 2.5 1.8 -28 10

Total Inorganic Nitrogenc

Northeastb 6.5 4.4 -32 3

Mid-Atlantic 8.7 6.0 -31 11

Southeastb 5.9 4.8 -19 2

Midwest 9.0 6.9 -23 10

aAll values are statistically significant at the 95% confidence level unless otherwise indicated.

bPercentage change in this region was not tested for statistical significance because too few monitoring sites are available.

c Total deposition is estimated from raw measurement data, not rounded, and may not equal the sum of dry and wet deposition.

Source: U.S. EPA, 2010c

Chapter 2 | Results of the Acid Rain Program

30
One recent scientific advancement pertinent to 

obtaining more accurate estimates of atmospheric 

deposition is the increasing use of statistical and 

empirical models supported by detailed spatial 

measurements that more accurately incorporate 

the enhancement of precipitation by orographic 

lift (i.e., where moist air is forced to a higher 

elevation by mountainous or other rising terrain, 

causing fog and/or precipitation) and other factors 

(Nanus et al., 2003; Grimm and Lynch, 2004; 

Weathers et al., 2006). Using such approaches to 

obtain better estimates of atmospheric deposition 

is important because most deposition monitoring 

sites tend to be located at the lowest elevations 

within sensitive ecosystems and, therefore, tend 

to underestimate deposition relative to a model 

that incorporates terrain and mountain effects into 

deposition estimates. For example, Weathers et 

al. (2006) estimated that total nitrogen and sulfur 

deposition levels for the Acadia National Park in 

Maine and the Great Smoky Mountain National 

Park in Tennessee and North Carolina were about 

70% greater than indicated by data collected 

from nearby wet and dry monitoring stations. 

These modeling results highlight that atmospheric 

deposition to sensitive high-elevation ecosystems 

is generally greater than indicated by data from 

most precipitation sampling sites, which tend to be 

in easily accessible locations at lower elevations.

Sulfur Deposition

Between the 1989 to 1991 and 2007 to 2009 

observation periods, decreases in wet deposition 

of sulfate averaged more than 40% for the eastern 

United States. Some of the greatest decreases 

have occurred in the mid-Appalachian region, 

including Maryland, New York, West Virginia, 

Virginia, and most of Pennsylvania (see Figure 

2-9). Less dramatic reductions have been observed 

across much of New England, portions of the 

southern Appalachian Mountains, and some areas 

of the Midwest. A principal reason for decreased 

sulfate deposition levels in the Northeast is a 

reduction in the long-range transport of sulfate 

from emission sources located in the Ohio River 

Valley. The decreases in sulfate documented in 

the Northeast, particularly across New England 

and portions of New York, were also affected by 

SO2 emission reductions in eastern Canada. 

Figure 2-9. Annual mean wet sulfate deposition in the United States for the 1989–1991 and 2007–2009 observation 

periods (U.S. EPA, 2010c).

0

4

8

12

16

20

24

28

> 32

1989 to1991 2007 to 2009

Note: Dots on all maps represent NADP monitoring sites. Data source: NADP

0

4

8

12

16

20

24

28

> 32

Wet SO4

2-

(kg/ha)

Wet SO4

2-

(kg/ha)

Portions of the eastern United States have also 

experienced significant decreases in dry and 

total sulfur deposition since the initiation of 

the ARP. Dry sulfur deposition levels in the 

mid-Atlantic region and the Midwest have 

decreased by 57% between the 1989 to 1991 

and 2007 to 2009 observation periods. Decreases 

in total sulfur deposition for the same areas 

were 50% and 53%, respectively, for the same 

period. Continuous data records for dry and 

31

NAPAP Report to Congress: An Integrated Assessment
total sulfur deposition are available from only 

a few sites in the Northeast and Southeast; 

therefore, the observed decreases in deposition 

may not be representative of these regions. 

Other studies reported in the scientific literature 

indicate decreases in sulfate concentrations in 

precipitation since the 1990s; these decreases are 

similar to those reported in this report. Historical 

data show that decreases in sulfate concentrations 

in precipitation are approaching 50% since the 

1980s for many monitoring sites in the eastern 

United States (Lehmann et al., 2005; Kvale 

and Pryor, 2006). Between the 1989 to 1991 

and 2007 to 2009 observation periods, sulfate 

concentrations in precipitation have decreased 

by approximately 40% throughout the eastern 

United States. A strong correlation between largescale SO2 emission reductions and large decreases 

in sulfate concentrations in precipitation has 

been noted in the Northeast and mid-Atlantic 

regions (Lehmann et al., 2005), which are among 

the areas most affected by acid deposition. 

Total Reactive Nitrogen

Historically, scientists have focused their studies of nitrogen deposition on the wet deposition of nitrate. Scientists recently expanded 

their focus to other facets of nitrogen deposition and increasingly assess nitrogen sources 

and deposition in terms of total reactive 

nitrogen. Total reactive nitrogen considers all 

biologically, chemically, and radiatively active nitrogen compounds in both wet and dry deposition, such as ammonia gas (NH3), ammonium 

ion (NH +

4 ), nitric oxide (NO), nitrite (NO –2 ), 

nitric acid (HNO3), nitrous oxide (N2O), 

nitrate (NO –

3 ), and organic compounds (e.g., 

urea, amines, nucleic acids). Sources of reactive 

nitrogen include agricultural practices, vehicle 

exhaust, and EGUs.

Chapter 2 | Results of the Acid Rain Program

32

Nitrogen Deposition (Nitrate 

and Ammonium)

Analyses of data from existing networks, such 

as NADP/NTN and CASTNET, indicate 

that nitrogen deposition has decreased since 

the early 1990s (see Figure 2-10); however, 

this decline was more significant after 2000. 

Decreases in nitrogen deposition in the United 

States have been less pronounced, show greater 

spatial variation in temporal trends, and are 

less certain than those of sulfur deposition for 

a variety of reasons, including the following:

•	 NOx emission declines have been less 

than the declines in SO2 emissions. 

•	 The	proportion	of	nationwide	NOx emissions 

that originates from EGUs is much lower 

(about 20%) than the proportion of SO2

emissions that originates from these same 

units (about 70%) (U.S. EPA, 2006a).

•	 NOx emissions originate from a number of 

non-ARP sources, including motor vehicles.

•	 The	contribution	of	ammonia	(NH3) emissions 

to overall nitrogen deposition originates largely 

from agricultural sources and is not regulated by 

the Clean Air Act (Stephen and Aneja, 2008). 

•	 A	large	number	of	chemical	species	of	nitrogen	

contribute to total nitrogen deposition, 

but are not well measured by existing 

monitoring networks (Sparks et al., 2008). 

Nitrogen deposition levels across the United States 

vary by region and by the type of deposition. 

Regional differences in these trends are apparent. 

The mid-Atlantic and northeastern states generally 

show the greatest downward trends in nitrogen 

deposition, and parts of the Southeast and West 

show the smallest decreases (Lehmann et al., 2005). 

In some regions (e.g., the Front Range of Colorado, 

Iowa), nitrogen deposition has increased in recent 

years (Burns, 2003; Ingersoll et al., 2008). Between 

the 1989 to 1991 and 2007 to 2009 observation 

periods, decreases in wet inorganic nitrogen 

(nitrate and ammonium) deposition levels in the 

eastern United States ranged from 16% in the 

Midwest to 27% in the mid-Atlantic region and 

the Northeast. Decreases in dry and total inorganic 
nitrogen deposition generally have been greater 

than those of wet deposition, with a 23% and 

31% decrease in total nitrogen deposition for the 

Midwest and the mid-Atlantic region, respectively.

Figure 2-10. Annual mean wet inorganic nitrogen deposition in the United States for the 1989–1991 and 2007–2009 

observation periods. (U.S. EPA, 2010c).

1989 to1991 2007 to 2009

Inorganic Inorganic Nitrogen Nitrogen (kg/ha) (kg/ha)

0.0 0.0

1.0 1.0

2.0 2.0

3.0 3.0

4.0 4.0

5.0 5.0

6.0 6.0

7.0 7.0

8.0 8.0

9.0 9.0

> 10.0 > 10.0

Note: Dots on all maps represent NADP monitoring sites. Data source: NADP

Figure 2-11. Stages of recovery for acidic lakes and streams (prepared by U.S. EPA).

Waters of Acute Concern

Water acidic all the time;

sensitive organisms

cannot survive

Waters of Low Concern

Recovery of water

chemistry; even sensitive

organisms can survive

Waters of Moderate Concern

Signicant recovery, but water

still acidied at times; sensitive

organisms may survive

Waters of Elevated Concern

Water still acidied seasonally

or after storms; sensitive

organism richness still low

Note: See Chapter 3 for further discussion of the ecological effects of acidification. 

Waters of acute concern are defined as having ANC less than 0 μeq/L. 

Waters of elevated concern are defined as having ANC from 0 to less than 50 μeq/L. 

Waters of moderate concern are defined as having ANC from 50 to less than 100 μeq/L. 

Waters of low concern are defined as having ANC greater than or equal to 100 μeq/L.

Surface Water Monitoring

Acid rain, resulting from SO2 and NOx emissions, 

is one of many large-scale anthropogenic effects 

that negatively impact the health of biota in lakes 

and streams in the United States. Surface water 

chemistry provides direct indicators of the potential 

effects of acidic deposition on the overall health 

of aquatic ecosystems. Since the implementation 

of the ARP, scientists have measured changes in 

some lakes and streams in the eastern United States 

and found signs of recovery in many, but not all, 

regions and waterbodies. Figure 2-11 provides an 

overview of the stages of recovery in acidic lakes 

and streams used in this report. As described in 

Chapter 3 of this report, organisms may survive in 

recovered lakes and streams, but may differ from the 

species that originally inhabited those waterbodies.

Three indicators of acidity in surface waters are 

used to provide information regarding both 

sensitivity to surface water acidification and the 

level of current and past acidification. These 

indicators are concentrations of sulfate (SO 2- 4 ) 

and nitrate (NO -

3 ) ions, and acid neutralizing 

capacity (ANC). Sulfate and nitrate are negatively 

charged ions with the potential to acidify waters 

and leach cations, including acidic aluminum 

cations, from watershed soils. Aluminum and H+

cations are known to be toxic to aquatic life, given 

high enough concentrations. Base cations, the 

by-products of weathering reactions that neutralize 

acids in watersheds, may also be indicators of 

changes in acidification. Assessments of acidic 

deposition effects, dating from the 1970s to the 

33

NAPAP Report to Congress: An Integrated Assessment

1989 to1991 2007 to 2009

Note: Dots on all maps represent NADP monitoring sites. Data source: NADP

0.0

1.0

2.0

3.0

4.0

5.0

6.0

7.0

8.0

9.0

> 10.0

0.0

1.0

2.0

3.0

4.0

5.0

6.0

7.0

8.0

9.0

> 10.0

Inorganic

Nitrogen

(kg/ha)

Inorganic

Nitrogen

(kg/ha)
present, have shown that sulfate is the primary 

negatively charged ion in most acid-sensitive waters 

(Driscoll et al., 2001). Sulfate ion concentrations 

in surface waters provide important information on 

the extent of base cation (i.e., calcium, magnesium, 

potassium, sodium) leaching in soils and provide 

insight on how sulfate concentrations relate to 

the levels of atmospheric sulfur concentrations 

and atmospheric sulfur deposition. Nitrogen is an 

important nutrient for plant growth; therefore, 

nitrogen inputs through deposition are incorporated

into forest biomass and soil organic matter, with 

only a fraction of the input leaching as nitrate 

into lakes and streams. However, as atmospheric 

nitrogen deposition increases to levels in excess 

of ecosystem needs, there is greater potential for 

increased leaching of nitrate into lakes and streams, 

leading to increased acidification. ANC is a measure 

of the acid-buffering capacity of water and an 

important indicator of the sensitivity and the degree

of surface water acidification or recovery that occurs 

over time. Acidification results in a diminished 

ability of water in a lake or stream to neutralize 

strong acids that enter aquatic ecosystems. 

Monitoring the trends in these indicators enables 

determination of whether conditions in acidsensitive lakes and streams are improving and 

heading toward chemical recovery or if conditions 

are degrading. Chemical recovery is a prerequisite 

for the subsequent recovery of aquatic plants and 

animals. Movement toward recovery of a lake 

or stream ecosystem is indicated by increases in 

ANC levels and decreases in sulfate and nitrate 

concentrations. Table 2-3 presents the percentage 

of monitoring sites in four acid-sensitive regions 

(i.e., New England, Adirondack Mountains, 

Northern Appalachian Plateau, and Central 

Appalachians) of the eastern United States (Figure 

2-12) with improving sulfate, nitrate, and ANC 

trends. The sites represent LTM lakes and streams 

monitored from 1990 to 2008. These regional 

trends were calculated using data from all the sites 

located within the region that had a complete 

data record for the time period considered. Trends 

are considered statistically significant if they 

exceed the 95% confidence interval (p < 0.05). 

As seen in Table 2-3 and Figure 2-13, significant 

improving trends in sulfate concentrations 

from 1990 to 2008 are found regionally and 

at nearly all monitoring sites in New England, 

the Adirondack Mountains, and the Northern 

Appalachian Plateau. However, in the central 

Appalachians, sulfate concentrations in 21% 

of individual streams monitored by the LTM 

program show degrading trends, and improving 

trends were noted at only 12% of sites. The central 

Appalachians have highly weathered soils that 

can store large amounts of deposited sulfate. As 

long-term sulfate deposition exhausts the soil’s 

ability to store sulfate, a decreasing proportion 

of the deposited sulfate is retained in the soil, 

and an increasing proportion is exported to 

surface waters. In addition, sulfate stored from 

decades of sulfur deposition leaches out from 

the soils over time. Thus, sulfate concentrations 

in streams in this region are increasing despite 

reduced levels of sulfate deposition.

Surface Water Monitoring 

Networks

Long-term surface water monitoring networks 

provide information on the chemistry of lakes 

and streams and on how waterbodies are 

responding to changes in emissions. Two EPAadministered monitoring programs provide 

information on the effects of acid rain on lakes 

and streams: the Temporally Integrated Monitoring of Ecosystems (TIME) program and the 

Long-Term Monitoring (LTM) program. These 

programs were designed to track the effect 

of the 1990 CAAA in reducing the acidity of 

lakes and streams in four acid-sensitive regions 

of the eastern United States. The surface 

water chemistry trend data in the four regions 

monitored by the TIME and LTM programs are 

essential for tracking the ecological response 

to ARP emission reductions. Additional information about these programs is available in 

Chapter 1.

Chapter 2 | Results of the Acid Rain Program

34
Adirondack Mountains

Northern Appalachian Plateau

Central Appalachians

LTM Monitoring Site

New England

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Figure 2-12. LTM site locations and regional groupings (prepared by U.S. EPA).

Table 2-3. Regional Trends in Sulfate, Nitrate, and ANC at LTM Sites from 1990 to 2008

Region Waterbodies 

Covered

% of Sites with 

Improving 

Sulfate Trenda

% of Sites with 

Improving 

Nitrate Trenda

% of Sites with

Improving 

ANC Trenda

Adirondack Mountains Lakes in NY 

N = 50

90 32 58

Northern Appalachian 

Plateau

Streams in NY, PA 

N = 9

78 33 56

New England Lakes in ME, VT 

N = 26

96 31 12

Central Appalachians Streams in VA

N = 60

12 45 12

a Trends determined by multivariate Mann-Kendall tests.

Data source: U.S. EPA

35

NAPAP Report to Congress: An Integrated Assessment
-2.5 -2 -1.5 -1 -0.5 0 0.5 1

Base Cations

Sulfate

Nitrate

ANC

Slope of Trend

New England lakes (n=26)

Adirondack Mountain lakes (n=50)

Northern Appalachian Plateau streams (n=9)

Central Appalachian streams (n=60)

Figure 2-13. Trends in acidification of LTM lakes and streams, 1990–2008 (prepared by U.S. EPA).

Chapter 2 | Results of the Acid Rain Program

36

Improving trends are increases in ANC levels 

and decreases in sulfate and nitrate concentrations. Degrading trends show decreases in 

ANC levels and increases in sulfate and nitrate 

concentrations.

Trends in surface water nitrate concentrations 

are also mixed (Figure 2-13) and indicate flat or 

slightly degrading nitrate trends at several individual 

LTM lake and stream sites (Table 2-3). Improving 

trends for nitrate concentrations were noted at 

only 31% to 45% of sites in the four regions. 

These trends do not appear to reflect changes in 

emissions or deposition in these areas and are likely 

a result of ecosystem factors such as soil retention, 

denitrification rates, and vegetation disturbance.

Declines in sulfate deposition levels likely result 

in many of the improving trends for ANC, but 

ANC levels still lag behind improving sulfate 

trends at many of the monitoring sites. From 

1990 to 2008, monitoring sites that showed a 

majority of improving ANC trends were in the 

Adirondack Mountains (58%) and Northern 

Appalachian Plateau (56%). Streams in the central 

Appalachians and in New England showed only 

a few statistically significant trends of improving 

ANC (12%) (Table 2-3). The declining trends of 

base cations (Figure 2-13) in lakes and streams may 

delay the onset of recovery. Decreasing base cation 

levels can balance out declining levels of sulfate 

and nitrate, preventing ANC from increasing. 

The observed improving trends for ANC can, in 

part, be attributed to declines in sulfate deposition 

levels. From 1990 to 2008, many of the statistically 

significant trends observed at individual LTM sites 

showed increasing ANC levels (improving trend) 

in surface waters. Overall, only seven sites indicate 

a statistically significant degrading trend in ANC.

In order to assess the impacts of the implementation 

of the ARP, the average ANC levels of lakes in the 

Northeast monitored under the TIME program 

were evaluated and compared for the 1991 to 

1994 and 2006 to 2008 time periods (see Figure 

2-14). Seven and a half percent of lakes in the 

1991 to 1994 time frame had 3-year mean ANC 

levels below 0 microequivalents per liter (μeq/L) 

and were categorized as being of acute concern 

(see text box). The percentage of lakes in this 

category dropped to 4.3% in the 2006 to 2008 time 

frame. Additionally, the percentage of lakes in the 

elevated concern category dropped from 13.8% 

for the 1991 to 1994 time period to 10.1% from 
2006 to 2008, while the percentage of lakes in the 

moderate concern category increased from 6.4% t

13.5%. These results point to a decrease in acidity, 

particularly for the subset of lakes with low ANC.

Case Study—Critical Loads 

for Lakes and Streams

The critical loads approach (see text box) is an 

assessment tool that can be used to determine the 

degree to which air pollution may be affecting 

ecological health. This approach provides a 

useful lens through which to assess the results of 

o 

current policies and programs and to evaluate 

the potential value of proposed policy options in 

terms of ecosystem protection. The critical loads 

approach has been employed routinely as an 

assessment tool for many years in the countries 

of the European Union and Canada. This case 

study will explore the extent to which lakes in th

Adirondack Mountains and streams in the centra

Appalachians are protected from acidifying nitro

and sulfur deposition as a result of Title IV emiss

reductions. Further discussion of critical loads 

can also be found in Chapter 3 of this report.

e 

l 

gen 

ion 

Acid Neutralizing Capacity

Lakes and streams with ANC values defined as 

less than or equal to 0 μeq/L are acidic and are 

considered to be of “acute concern.” Lakes and 

streams having springtime ANC values less than 

50 μeq/L are generally considered “sensitive” to 

acidification and are of “elevated concern.” Lakes 

and streams with ANC higher than 50 μeq/L are 

generally considered less sensitive or insensitive 

to acidification and may be considered of “moderate concern” (i.e., ANC levels between 50 and 

100 μeq/L) or “low concern” (i.e., ANC levels 

greater than 100 μeq/L). When ANC is low, and 

especially when it is negative, stream water pH is 

also low (certainly less than 6 and commonly less 

than 5), and there may be adverse impacts on fish 

and other animals essential for a healthy aquatic 

ecosystem (see table below).

Aquatic Ecosystem Concern Levels and Ecological Effects

Category 

Label

ANC Level 

(μeq/L)

Expected Ecological Effects

Acute 

Concern

Less than 0 Near-complete loss of fish populations is expected. Planktonic communities 

have extremely low diversity and are dominated by acid-tolerant forms. The 

numbers of individuals in plankton species that are present are greatly 

reduced.

Elevated 

Concern

0 to less 

than 50

Fish species richness is greatly reduced (e.g., more than half of expected 

species are missing). On average, brook trout populations experience 

sublethal effects, including loss of health and reproduction (fitness). During 

episodes of high acid deposition, brook trout populations may experience 

lethal effects. Diversity and distribution of zooplankton communities decline.

Moderate 

Concern

50 to less 

than 100

Fish species richness begins to decline (e.g., sensitive species are lost from 

lakes). Brook trout populations are sensitive and variable, with possible 

sublethal effects. Diversity and distribution of zooplankton communities 

begin to decline as species that are sensitive to acid deposition are affected.

Low 

Concern

Greater than 

or equal to 100

Fish species richness may be unaffected. Reproducing brook trout populations are expected where habitat is suitable. Zooplankton communities are 

unaffected and exhibit expected diversity and distribution.

37

NAPAP Report to Congress: An Integrated Assessment
80

70

60

ese of Lak

50

g

40

centa

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erP

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10

0

Acute Elevated Moderate Low

Concern Concern Concern Concern

1991–1994 2006–2008

Figure 2-14. Northeastern TIME Lakes by ANC 

Category, 1991-1994, vs. 2006-2008. Acute Concern—

less than 0 μeq/L, Elevated Concern—0 μeq/L to less 

than 50 μeq/L, Moderate Concern—50 μeq/L to less 

than 100 μeq/L, Low Concern—greater or equal to 

100 μeq/L (prepared by U.S. EPA).

Chapter 2 | Results of the Acid Rain Program

38

Critical Loads

A critical load is a quantitative estimate of 

exposure to one or more pollutants, below 

which significant harmful effects on specific 

sensitive elements of the environment do not 

occur according to present knowledge (Nilsson and Grennfelt,1988). If pollutant exposure 

is less than the critical load, adverse ecological 

effects (e.g., reduced reproductive success, 

stunted growth, loss of biological diversity) 

are not anticipated, and recovery is expected 

over time if an ecosystem has been damaged 

by past exposure. A critical load exceedance is the measure of pollutant exposure 

above the critical load. This means the pollutant exposure is higher than, or “exceeds,” the 

critical load and the ecosystem continues to 

be exposed to damaging levels of pollutants. 

The critical load for a lake or stream provides a 

benchmark against which to assess the extent that 

a waterbody is potentially at risk due to current 

deposition levels. The analysis focuses on the 

combined load of sulfur and nitrogen deposition 

below which the ANC level would still support 

healthy aquatic ecosystems. There are numerous 

methods and models that can be used to calculate 

critical loads for acidity. Drawing on the peerreviewed scientific literature (e.g., Dupont et 

al., 2005), both case studies use the Steady-State 

Water Chemistry (SSWC) model to calculate 

the critical load. Critical loads were calculated 

for 117 lakes in the Adirondack Mountains and 

184 streams in the central Appalachian region. 

The modeled lakes and streams in both case 

studies are a statistical population of waterbodies 

that represents 1,830 lakes of various sizes found 

throughout the Adirondack Mountains with 

surface areas ≥ 4 ha and 70,000 km of streams 

in the central Appalachians. Both statistical 

populations were determined by EMAP. 

For these particular analyses, the critical load 

represents the combined deposition loads of sulfur 

and nitrogen to which a lake or stream could be 

subjected and still have an ANC of 50 μeq/L or 

higher. While a critical load can be calculated 

for any ANC level, 50 μeq/L was chosen because 

it provides a moderate level of protection for 

most fish and other aquatic organisms, although 

systems can still become episodically acidic and 

some sensitive species may be lost. Critical loads 

of combined total sulfur and nitrogen deposition 

are expressed in terms of ionic charge balance as 

milliequivalents per square meter per year (meq/m2/

yr). When actual measured deposition of nitrogen 

and sulfur is greater than the critical load, the 

critical load is “exceeded,” meaning that combined 

sulfur and nitrogen deposition was greater than a 

lake or stream could sustain and still maintain the 

ANC level of 50 μeq/L or above. In some cases, 

negative critical loads may occur when the supply 

of base cations is naturally low. In these cases, 

the lake or stream would have a natural, pre-
Adirondack mountain lakes are sensitive to acidic deposition (photo courtesy of NASA).

39

NAPAP Report to Congress: An Integrated Assessment

acidification ANC level below 50 μeq/L. An ANC 

of 50 μeq/L or greater would not be achievable 

regardless of changes in acidic deposition levels. 

In order to assess the extent to which regional 

lake and stream ecosystems are protected by the 

emission reductions achieved by Title IV, these 

case studies compare the amount of deposition 

systems can receive—the critical load—to measured

deposition for the period before implementation 

of the ARP (1989–1991) and for a recent period 

after ARP implementation (2006–2008).

Overall, these critical load analyses show that 

emission reductions achieved by the ARP have 

resulted in improved environmental conditions 

and increased ecosystem protection in both the 

Adirondack Mountains and the central Appalachian

region. However, the analyses also show that, 

despite some ecological recovery, lakes and streams 

in these regions remain at risk due to current acid 

deposition levels. Thus, further emission reductions 

are necessary for recovery of these sensitive aquatic 

systems and ecosystem protection. In Chapter 4 of 

this report, a modeling analysis examines the degree

of ecosystem recovery and protection that would 

be achieved through further emission reductions.

Adirondack Mountain Lakes

The Adirondack Mountains are located in northern 

New York. Many lakes in this region are sensitive 

to acidic deposition because of the abundance of 

shallow soils with low levels of base cations (i.e., 

calcium, magnesium, sodium, potassium) and the 

underlying igneous bedrock with low weathering 

rates and limited ability to buffer acidity (Driscoll 

et al., 1991; Sullivan et al., 2006a). Decreases in 

regional sulfur deposition have resulted in decreased 

sulfate concentrations in Adirondack Mountain 

lakes, which is an important sign of ecological 

recovery (see Figure 2-13). While inter-annual 

variability in ANC levels and nitrate concentrations 

is evident in this region, the overall trend in 

these lakes is a slight increase in ANC levels and 

modestly decreasing nitrate concentrations.

On average, the calculated critical load of sulfur and 

nitrogen for lakes in the Adirondack Mountains 

is 162 meq/m2/yr. For the period from 1989 

to 1991, before ARP implementation, 45% of 

the lakes examined received levels of combined 

sulfur and nitrogen deposition that exceeded the 

critical load (Figure 2-15). For the 2006 to 2008 

period, 30% of lakes exceeded the critical load 

(Figure 2-16). Thus, during the 2006 to 2008 

period, 15% fewer of the Adirondack Mountain 

lakes examined were receiving sulfur and nitrogen 

deposition loads that threaten the health of these 

ecosystems. Although lakes in the Adirondack 

Mountains have improved, over 550 lakes still 

exceed their calculated critical load of sulfur and 

nitrogen based on the statistically representative 

population of lakes determined by EMAP.
 

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Adirondack Park Boundary

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Figure 2-15. Critical load exceedances in Adirondack Mountain TIME lakes, 1989 to 1991 (No. lakes = 117) 

(prepared by U.S. EPA).

Chapter 2 | Results of the Acid Rain Program

40

 

 

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Figure 2-16. Critical load exceedances in Adirondack Mountain TIME lakes, 2006 to 2008 (No. lakes = 117) 

(prepared by U.S. EPA).
Central Appalachian Streams

The central Appalachian region of Virginia and 

West Virginia, including Shenandoah National 

Park, is known to be sensitive to acidic deposition. 

Base-poor soils and low weathering rates of the 

bedrock that underlies this region result in streams 

that are particularly susceptible to acidification. 

As a result, many miles of streams in the region 

have been impacted by acid deposition, with 

greatly reduced diversity of aquatic life, including 

important recreational fish species such as brook 

trout. The declines in deposition of sulfur and 

nitrogen since the 1990s have resulted in only 

slight signs of recovery from acid rain in the streams 

of this region. Figure 2-13 shows monitored 

trends in sulfate and nitrate concentrations and 

ANC levels for central Appalachian streams. 

Sulfate concentrations in these streams have 

remained level, while nitrate concentrations 

and ANC levels have improved slightly.

On average, the calculated critical load of sulfur 

and nitrogen for streams in the central Appalachian 

region is 370 meq/m2/yr. For the period before 

ARP implementation (1989 to 1991), 41% 

of streams in the central Appalachian region 

received levels of combined sulfur and nitrogen 

deposition that exceeded the critical load, and 

therefore, could not be adequately neutralized by 

the environment (Figure 2-17). For the period 

from 2006 to 2008, 31% of the streams examined 

continued to receive acid deposition levels that 

exceeded the critical load (Figure 2-18). Thus, 

only an additional 10% of streams in the central 

Appalachian region are protected from ecological 

damage from acidic deposition, as compared to 

the period before ARP implementation. Although 

some streams in the central Appalachians likely have 

improved, approximately 35,000 km of streams still 

exceed their calculated critical load of sulfur and 

nitrogen based on the statistically representative 

population of lakes determined by EMAP.

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Critical Load Exceedances (1989–1991)

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Figure 2-17. Critical load exceedances in central Appalachian streams, 1989 to 1991 (No. streams = 184) 

(prepared by U.S. EPA).

41

NAPAP Report to Congress: An Integrated Assessment
Visibility 

In the atmosphere, SO2 and NOx gases are 

transformed into fine particles of sulfate and 

nitrate. Sulfate and nitrate particles scatter and 

absorb light energy, impairing visibility and 

contributing to regional haze. Visibility refers to 

the clarity with which distant objects are perceived 

(U.S. EPA, 2001). Sulfate is generally the dominant 

contributor to visibility impairment in the eastern 

United States and a large contributor in the western 

United States. Nitrate is a significant contributor 

to visibility impairment in parts of California and 

in the Midwest during the winter. Changes in 

visibility conditions are measured in deciview (dv) 

units, which are approximately proportional to 

the perceived change in haze levels and increase as 

visibility decreases. A one dv change is a small but 

noticeable change under many viewing conditions. 

Average, naturally occurring conditions (without 

pollution) are about 7 dv to 8 dv in the eastern 

United States and 4.5 dv to 5.5 dv in the western 

United States. For the worst haze days, visibility 

conditions exceed natural levels by 10 dv to 18 

dv in the eastern United States and by about 

3 dv to 15 dv in the western United States.

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Figure 2-18. Critical load exceedances in central Appalachian streams, 2006 to 2008 (No. streams = 184) 

(prepared by U.S. EPA).

Visibility Measures 

Visibility is measured in “deciviews”, which is a 

way of characterizing the relationship between 

air pollution and perception of visibility. The 

clearest days are those that demonstrate the 

highest daily visibility (i.e., lowest deciview [dv] 

values) for a given year. Trends among the clearest days are assessed by comparing the clearest 

20% of days every year (i.e., best 20% of visibility). 

The haziest days are those that demonstrate the 

lowest daily visibility (i.e., highest dv values) for 

a given year. Trends among the haziest days are 

assessed by comparing the haziest 20% of days 

every year (i.e., worst 20% of visibility) 

(U.S. EPA, 2003).

Chapter 2 | Results of the Acid Rain Program

42
Data from the IMPROVE network, which monitors 

visibility in and near national parks and wilderness 

areas, show no statistically significant trends in 

visibility at most of the 128 sites monitored over 

the 10-year period from 1999 to 2008 (see Figures 

2-19 and 2-20). Of the sites with statistically 

significant trends for either the best or worst 20% 

of visibility conditions, most have improving 

visibility. From 1999 through 2008, for the best 

visibility conditions, 45 sites show significant 

improvements and no sites show significant decline; 

for the worst visibility conditions, 10 sites show 

significant improvement and 5 sites show significant 

decline (all trends are at the p < 0.05 significance 

level).The Northeast has sites with improving 

visibility on both best and worst visibility days, 

principally due to regionally decreased sulfate 

particulate concentrations. Improvement for these 

northeastern sites is about 5 dv for worst days 

and 3 dv for best days over the 10-year period. 

In the Pacific Northwest, a significant improving 

trend of about 3 dv to 5 dv per decade for worst 

haze periods is found at monitoring sites near 

and/or downwind from Seattle and Portland and 

may reflect changing impacts of these urban/

industrial areas. A better understanding of the 

causes of individual site trends requires additional 

analysis. Trends for each individual site shown on 

the map in Figures 2-19 and 2-20 can be viewed 

at http://views.cira.colostate.edu/web/Trends. 

20% Clearest Days – dv

Trends Improving

 No Change

 Trends Deteriorating

0 125 250 500

Kilometers

Note: Trends were determined by a Theil 

trend test at the p < 0.05 signicance level.

Figure 2-19. 1999 to 2008 trends for the change in visibility on the clearest days (best 20% of visibility) as 

measured at 140 IMPROVE monitoring stations (prepared by NPS).

As dv measurements increase, visibility 

decreases.

43

NAPAP Report to Congress: An Integrated Assessment
0 2 125 50 500

Kilome ters

20% Haziest Days – dv

Trends Improving

 No Change

 Trends Deteriorating Note: Trends were determined by a Theil 

trend test at the p < 0.05 signicance level.

Figure 2-20. 1999 to 2008 trends for the change in visibility on the haziest days (worst 20% of visibility) as 

measured at 140 IMPROVE monitoring stations (prepared by NPS).

Chapter 2 | Results of the Acid Rain Program

44
CHAPTER 3

Effects of Acid Deposition 

on Ecosystems: Advances in 

the State of the Science

State of the Science at a Glance

•	 Scientists	have	observed	delays	in	

ecosystem recovery in the eastern United 

States, despite decreases in emissions and 

deposition over the past 30 years.

•	 A	growing	body	of	literature	documents	

impacts of nitrogen deposition on western 

ecosystems.

•	 “Critical	loads”	are	increasingly	used	by	

scientists as a tool for quantifying the sulfur 

and nitrogen deposition levels at which 

ecosystems are impacted in order to better 

inform air quality policies.

•	 Recent	literature	identifies	linkages	

between a changing climate, an altered 

carbon cycle, and ecosystem response to 

acid deposition as an important emerging 

area	of	scientific	investigation.

•	 Scientists	identify	multi-pollutant	

interactions, including reactions among SO2, 

NOx,	ozone,	and	mercury,	as	an	understudied area of ecosystem impacts, which 

if better understood, would likely have 

implications for future air pollutant policies.

Chapter 2 focused on the environmental 

results of the ARP, presenting data from 

national monitoring networks on SO2 and 

NOx emissions, air quality, atmospheric 

deposition, surface water chemistry, 

and visibility. This chapter expands on 

this information by examining the most 

recent research into how ecosystems 

respond to acid deposition, especially the 

processes that control the recovery of 

ecosystems as acid deposition decreases.

In Chapter 2, two general trends were discussed 

regarding the current recovery status of affected 

ecosystems: (1) these ecosystems are trending 

generally towards recovery, but improvements 

in ecosystem condition shown by surface water 

chemistry monitoring data thus far have been 

less than the improvements in deposition; and 

(2) ecosystem impacts and trends vary widely 

by geographic region, but the evidence of 

improvement is strongest and most evident in the 

Northeast. These trends are not uniform across 

the United States, however, and in some regions 

(e.g., central Appalachian Mountain region), trends

in improved water quality are generally not evident.

Despite the strong link in many areas between 

reduced emissions and reduced acidity of 

atmospheric deposition, the link is less clear 

between reduced acidity and recovery of the 

biological communities that live in aquatic and 

terrestrial ecosystems that have experienced 
Chapter 3 | Effects of Acid Deposition on Ecosystems

46

deleterious effects from acid deposition. The 

recovery of these communities is proceeding at a 

slower pace than, for example, the improvements in

stream and lake ANC would indicate. The goal of 

this chapter is to synthesize the science in a weightof-evidence manner to provide policy makers 

with tangible evidence and likely causative factors 

regarding ecosystem status and recovery patterns to 

date. This chapter serves as an update to the 2005 

NAPAP RTC (NSTC, 2005), with an emphasis 

on scientific studies and monitoring since 2003, 

which was the last year for consideration of research

results in the 2005 report. Several issues pertinent 

to ecosystem response to emission controls and 

acid deposition are receiving increasing attention 

in the scientific literature and will be discussed in 

this chapter, including the (1) observed delay in 

ecosystem recovery in the eastern United States, 

even with decreases in emissions and deposition 

over the past 30 years; (2) emerging ecosystem 

impacts of nitrogen deposition in the western 

United States; (3) the application of critical 

deposition loads as a tool for scientists to better 

inform air quality policies; (4) the role of changes 

in climate and the carbon cycle as factors that affect

the response of ecosystems to acid deposition; 

and (5) the interaction of multiple pollutants in 

ecosystems. Throughout this chapter, the value 

of long-term environmental monitoring data in 

informing air quality policy will be highlighted, 

including the limitations of assessing the current 

status of some ecosystem indicators for which 

continuous, long-term data are lacking. 

Effects of Atmospheric 

Deposition on Aquatic and 

Terrestrial Ecosystems

This section expands on the surface water 

chemistry section of Chapter 2 and describes 

the latest studies on the spatial and temporal 

patterns of recovery in a variety of ecosystems 

that are sensitive to acid deposition. Factors that 

control ecosystem sensitivity to deposition are also 

discussed, with an emphasis on the links between 

terrestrial and aquatic ecosystems. This section 

first considers the eastern United States, then the 

Rocky Mountain region, the far western United 

States, and finally the coastal areas of the United 

States. This is a logical division because the effects 

of atmospheric deposition and issues discussed 

by scientists and policy makers are somewhat 

different across these regions and ecosystems.

Ecosystem Recovery in Aquatic 

and Terrestrial Ecosystems of 

the Eastern United States

The regions of the United States with an abundance 

of acid-sensitive waters and ecosystems were 

well recognized by the end of the 1980s during 

the decade of intensive research performed 

under the original NAPAP. These acid-sensitive 

ecosystems are generally located in upland, 

mountainous terrain in the eastern United States 

and are underlain by bedrock that is resistant 

to weathering, such as granite or quartzite 

sandstone. The following sections examine trends 

in surface water chemistry; the impacts of episodic 

acidification on surface waters; the effects of stored 

sulfate in the soils of the Southeast; base cation 

depletion in soils; and how aquatic and terrestrial 

ecosystems recover from acidifying deposition.

Nitrate Concentration Trends

Trends in nitrate concentrations in headwater 

surface waters are influenced by natural disturbances such as insect defoliation and ice storms, 

as well as human disturbances such as forest 

harvesting, and are therefore not as clearly linked 

to trends in nitrogen deposition as are those of 

sulfur. Nonetheless, over the long term, decreases 

in NOx emissions from the power generation 

sector are expected to assist in efforts to reduce 

nitrate concentrations in surface waters.

Patterns in Surface Water 

Chemistry and Soils

Monitoring of lakes and stream water chemistry 

in the eastern United States has shown the 

continuation of widespread water quality 

improvements that are consistent with a trend 

towards recovery in these aquatic ecosystems (Kahl 

et al., 2004; Driscoll et al., 2007). Decreasing 
NAPAP Report to Congress: An Integrated Assessment

47

trends in sulfate concentrations have been the 

sharpest, generally showing decreases of about 

2 to 3 μeq/L/yr since the early 1990s in waters 

in the Northeast and the Northern Appalachian 

Plateau. These trends are consistent with the large 

decreases in SO2 emissions and sulfur deposition 

and are paralleled by some improvement in 

pH, ANC, and aluminum (i.e., Al2+ and Al3+) 

concentrations. However, the improvements in 

pH, ANC, and aluminum concentrations have 

generally been much smaller than the decreases 

in sulfate concentrations. Nitrate concentrations 

in these surface waters have also shown smaller 

decreases that have been less persistent and less 

consistent than those of sulfate concentrations 

(Kahl et al., 2004; Burns et al., 2005; Driscoll et 

al., 2007). The most recent trends in surface water 

chemistry from the EPA LTM monitoring program 

are reported in Chapter 2 and shown in Table 2-3. 

These trends are generally consistent with those 

reported in the studies cited here, indicating sharp 

decreases in sulfate concentrations, some modest 

decreases in nitrate concentrations, and increases 

in ANC in the eastern United States. Streams in 

the central Appalachian region show the weakest 

trends in recovery. At least over periods of a decade 

or more, trends in nitrate concentrations in surface 

waters are not as clearly linked to trends in nitrogen 

deposition as trends in sulfate concentrations 

in surface waters are linked to trends in sulfur 

deposition (Aber et al., 2002; Burns et al., 2006; 

Eshleman et al., 2009). Nonetheless, if temporal 

trends follow observed spatial patterns of nitrogen 

deposition and surface water nitrate concentrations 

over the long term, then decreases in these nitrate 

concentrations are expected to eventually follow 

decreases in nitrogen deposition (Aber et al., 2003).

Aluminum (i.e., Al2+ and Al3+) levels in surface 

water are monitored because acidifying 

deposition releases aluminum from soils and 

high concentrations of aluminum are toxic to 

many aquatic biota.

The Laurel Prong fork of the Rapidan River in Virginia 

(photo courtesy of U.S. EPA).

Surface water chemistry monitoring programs 

that document the effects of acid deposition 

tend to collect samples at regular time intervals, 

resulting in data that primarily reflect baseflow 

or non-storm conditions (although there are 

exceptions). The shortcoming of such a sampling 

strategy is that surface waters, especially streams 

and small, shallow lakes, tend to acidify during 

rainstorms and snowmelt. This phenomenon is 

termed episodic acidification and has been well 

documented (Lawrence, 2002; Deviney et al., 

2006). Additionally, episodic acidification has been 

shown to affect surface water biota and to slow 

ecosystem recovery in waters that otherwise appear 

to be improving based on data collected at low-flow 

conditions (Passy et al., 2006; Kowalik et al., 2007).

These findings suggest that biological recovery 

will hinge in part on the acidity of surface waters 

declining to a level at which minimal episodic 

acidification occurs because the survival of acidsensitive aquatic species is dependent on both the 

magnitude and duration of episodic acidification 

(Baldigo and Murdoch, 1997). Although lowflow ANC values in the range of 30 to 50 μeq/L 

have been suggested to provide protection against 

acidification and biological harm during high-flow 

periods (Driscoll et al., 2001; Kahl et al., 2004), 

other studies have shown that episodic acidification 

occurs in streams with low-flow ANC values as 

high as 100 to 150 μeq/L (Lawrence, 2002). Some 

evaluations of ecosystem recovery (e.g., Kahl et 

al., 2004) provide percent-improvement values for 

various metrics of stream acidity and the extent 

of acidic waters in a region. These evaluations are 

based largely on low-flow conditions or chronic 

acidification scenarios and do not fully consider 

episodic acidification. One study in Maryland 
Chapter 3 | Effects of Acid Deposition on Ecosystems

48

evaluated changes in episodic stream acidification 

from the early 1990s to the late 1990s and found 

little change in the magnitude of acidification 

during storms. However, the study also found a shift

towards a diminished role for sulfate in episodic 

acidification, suggesting a decreasing contribution 

of acid deposition to this phenomenon in parallel 

with declining sulfate deposition (Kline et al., 

2007). Studies by Rice et al. (2006) and Lawrence 

et al. (2008) compared the percent of streams or 

watershed area that is chronically acidified with 

the percent that is episodically acidified in the 

Shenandoah National Park, VA, and the western 

Adirondack Mountains, NY, respectively. In each 

case, the total percentage of acidified streams or 

watershed area increased substantially when episodic 

acidification was considered (see Table 3-1).

One noteworthy exception to the generalizations 

above about geographic patterns of improvement 

in surface water chemistry is watersheds in the 

Southeast (e.g., the central Appalachian region of 

Virginia) where stream ANC values have generally 

not improved and sulfate concentrations have show

little change since the early 1990s (Kahl et al., 2004

Webb et al., 2004; U.S. EPA, 2009a; Chapter 2 of 

this report). The lack of chemical recovery in the 

streams of the central Appalachian region is believed

to be due in part to the old and highly weathered 

condition of non-glaciated southern soils. These 

soils contain an abundance of adsorbed sulfate that 

can be slowly released as sulfuric acid deposition 

decreases, slowing ecosystem recovery (Webb et 

n 

; 

al., 2004). In these soils, decades of sulfate release 

and decreased deposition levels may be necessary 

to shrink the pool of sulfate in the soils to the 

point where acidification of streams no longer 

occurs. Ecosystem recovery of acidified watersheds 

in the highly weathered soils of the Southeast will 

likely continue to lag behind that of the glaciated 

Northeast by several years to several decades.

Table 3-1. Comparison of Chronically and Episodically Acidified Streams in a Watershed

Location Percent of Streams 

or Watershed 

Area That is 

Chronically Acidic

Percent of Streams Total Percent of Streams Reference

or Watershed or Watershed Area 

Area That is That is Chronically and 

Episodically Acidic Episodically Acidic

Shenandoah 8.7 22.5 31.2 Rice et al., 

National Park, VAa 2006

Western Adirondack 

Mountains, NYb

35 30 65 Lawrence et al., 

2008

a

b Chronic acidification is defined as percent of streams with a base cation surplus (a surrogate for ANC that eliminates the 

contribution from dissolved organic carbon [DOC]) value < 25 μeq/L during summer, and episodic acidification is defined by 

the same threshold value during snowmelt conditions.

Chronic acidification is defined as the percent of watershed area with ANC < 0 μeq/L for at least 168 successive hours at a 

4-year recurrence interval, and episodic acidification is defined as the percent of watershed area with ANC < 20 μeq/L for at 

least 72 successive hours at a 4-year recurrence interval.

Ecosystem	recovery	of	acidified	watersheds	

in highly weathered soils of the Southeast 

will likely continue to lag behind that of the 

Northeast by several years or, more likely, 

decades.

Recent research has shown that decreases in base 

cations (e.g., calcium, magnesium, potassium, 

sodium) in soils may also be slowing the recovery 

of aquatic and terrestrial ecosystems even though 

they are receiving reduced loads of acid deposition 

(Bailey et al., 2005; Sullivan et al., 2006b; Johnson 

et al., 2008; Warby et al., 2009). The base cations 

calcium, magnesium, and potassium are important 

nutrients for the growth and health of forest 

vegetation and are readily accessed from the pool of 

cations held on the surfaces of soil particles. These 

base cations can be displaced from the surfaces 

of soil particles by hydrogen and aluminum ions 

and leached into local waters, effectively removing 

these nutrients from soils and lowering the levels 

that are available to the plants growing in the 
NAPAP Report to Congress: An Integrated Assessment

49

soils. Acid deposition accelerates this as part of 

the acidification process, which initially leads to 

increases in base cation concentrations in surface 

waters, but eventually acidifies the waters as well. 

Base cation depletion refers to this accelerated loss 

of these cations from soils. Depletion of calcium 

is of particular concern because this cation is in 

high demand by sugar maple (St. Clair et al., 

2008). Although a variety of mechanisms (e.g., 

maturation of forest vegetation) contribute to 

base cation depletion, acid deposition and the 

consequent mobilization of aluminum in soils have 

been identified as important drivers of depletion 

(Lawrence et al., 1995). Acid deposition and 

aluminum mobilization have contributed to dieback 

and decline of sugar maple (Bailey et al., 2004; 

Hallet et al., 2006; St. Clair et al., 2008) and to 

freezing injury in red spruce (DeHayes et al., 1999; 

Hawley et al., 2006). Recent work has found that 

declining growth rates can begin in apparently 

healthy sugar maple trees up to two decades before 

a strong decline in tree health becomes visually 

evident (Long et al., 2009). Other studies have 

found linkages between acid deposition, soil calcium 

availability, and the reproductive success of bird 

species (Graveland, 1998; Hames et al., 2002).

Sugar maple showing brilliant fall colors (photo courtesy 

of Jacquie Tinkler, NPS).

Base	cation	depletion	is	a	high-priority	area	

for	future	acidification	research	and	is	of	

high relevance to the development of future 

air quality policies designed to protect 

ecosystems.

The ANC value of a waterbody reflects the sum 

of the base cation concentrations (calcium, 

magnesium, potassium, sodium) minus the sum 

of the acid anion concentrations (primarily sulfate 

and nitrate). Depletion of base cations from soils 

over time will provide lower concentrations of 

these base cations to waters and result in a lower 

ANC value than would otherwise be predicted 

based on the acid anion concentrations alone. In 

this way, losses of exchangeable soil base cations 

can slow the recovery of aquatic ecosystems (Likens 

et al., 1996; Lawrence et al., 1999). The scientific 

consensus is that base cation depletion is slowing 

the rate of chemical recovery of lakes and streams 

and, therefore, also the rate of biological recovery 

in many of the sensitive aquatic ecosystems targeted 

by Title IV of the CAAA (Driscoll et al., 2001). 

Re-supply of these base cations generally occurs 

when chemical breakdown (i.e., weathering) of 

minerals in bedrock begins to release cations to 

soils and waters at a rate that is faster than the 

leaching loss caused by acid deposition. Acid 

deposition must likely decrease below a critical level 

that will vary among ecosystem locations before 

large increases in the ANC of acidified waters can 

occur. The role of base cation depletion is even 

greater when considering episodic acidification 

because soil chemistry plays an even greater role in 

stream chemistry during high-flow conditions than 

during low-flow conditions (Lawrence, 2002).

Important complications to this simplified 

conceptualization of the link between acid 

deposition and base cation depletion arise from 

the uptake of base cations by vegetation and from 

other factors. For example, the cation balance is also 

controlled in part by forest age, health, and rootmining ability, as well as harvesting practices (Blum 

et al., 2002; Hamburg et al., 2003; Grigal and 
Chapter 3 | Effects of Acid Deposition on Ecosystems

50

Ohmann, 2005). The exact rate of acid deposition 

that will allow recovery of base cation–depleted 

soils is not well known because the rate of mineral 

weathering is not well quantified in most soils, can 

be highly variable among locations, and is difficult 

to measure in the field (Bailey et al., 2003). Current 

models (such as the Model of Acidification of 

Groundwater in Catchments [MAGIC]) that are 

commonly applied to make quantitative predictions 

about acidification and critical loads in sensitive 

ecosystems can be, and have been, used to examine 

the role of base-cation depletion; however, they 

have not yet been applied widely for this purpose in 

the United States. Base-cation depletion is a highpriority area for future acidification research and is 

of great relevance to the development of future air 

quality policies designed to protect ecosystems.

Recovery of Biota

Aquatic Recovery

Surface water chemistry data commonly have been 

favored as an ecosystem indicator of acidification 

because these data are less expensive and more 

straightforward to obtain than data on aquatic 

biological populations. In addition, many studies 

have shown thresholds that link water chemistry 

values to the survival and reproductive success of 

organisms, such as brook trout, mayfly, and many 

others (Schofield, 1976; Baker and Christensen, 

1991). Using surface water chemistry data alone 

to assess ecosystem acidification assumes that 

when water chemistry degrades below a certain 

threshold, acid-sensitive organisms will be replaced 

by acid-tolerant organisms, and that when the 

water chemistry improves to a chemical threshold 

capable of successfully supporting more acidsensitive organisms, then the organisms will 

soon return to the waterbody. However, this is 

an oversimplified view of how ecosystems will 

respond to changes in lake or stream water quality. 

Acidification results in changes not only to the 

constituent or habitat that may be targeted as 

a threshold indicator but also to a wide variety 

of other chemical constituents and physical 

habitats. As a result, a complex set of changes in 

the ecological community may occur, including 

a shift in the members of the community from 

acid-sensitive to acid-tolerant species. Acidtolerant communities that develop in response to 

acidification may then resist change as conditions 

become less acidic, limiting the re-establishment 

of the acid-sensitive species that were originally 

impacted by acidification (Frost et al., 2006). Other 

factors may also slow this process of biological 

recovery, including physical limitations on dispersal, 

the proximity of species that may have historically 

resided in the water prior to acidification (Yan 

et al., 2003; Langford et al., 2009), and calcium 

concentrations that tend to be lower relative to 

ANC during ecosystem recovery than during 

acidification (Jeziorski et al., 2008; Cairns and 

Yan, 2009). These factors act to delay biological 

recovery relative to chemical recovery (Snucins, 

2003), and some scientists have suggested that 

active ecosystem restoration (e.g., reintroduction of 

previously lost species) may be needed to meet the 

goal of a “pre-acidification” biological community 

(Snucins and Gunn, 2003; Binks et al., 2005). 

There have been few studies in the United 

States on the recovery of aquatic biota over 

a	period	of	gradually	improving	acidification	

status; this is largely because none of the 

existing U.S. monitoring networks regularly 

provide such evaluations and because most 

available	data	originate	from	ad-hoc	studies.

Recreational fishing in South Rolly Lake in Alaska. (photo 

courtesy U.S. Fish and Wildlife Service). 
NAPAP Report to Congress: An Integrated Assessment

51

When a different pathway of change is followed 

during recovery than was followed during 

acidification, this behavior is called “hysteresis.” 

A conceptual representation of hypothetical 

acidification and recovery patterns in a waterbody 

is presented in Figure 3-1. Point 1 represents the 

pre-acidification status of water quality and the 

biological community. Point 2 shows the status 

when conditions were most acidic, and Point 3 

represents current conditions where hysteretic 

behavior is exhibited because the rate of recovery 

does not match the rate of acidification. At this 

point, the water quality has shown some recovery, 

but a pre-acidification biological community 

has not yet been achieved. The two dashed lines 

represent possible future recovery trajectories for 

the biological community as future acid deposition 

levels continue to decrease. These recovery 

patterns may or may not be linear as shown, and 

some evidence suggests that “threshold” behavior 

marked by a sharp change in slope with time 

may occur in biological communities recovering 

from acidification (Monteith et al., 2005). One 

trajectory ends at Point 4, where water quality has 

fully recovered, but the biological community has 

only recovered slightly. The other trajectory returns 

to Point 1, representing a full recovery of water 

quality and the biological community. Some studies 

indicate that even with full recovery of water quality 

(less acidic stream on the right side of the figure), 

the biological community will likely differ from 

what existed prior to acidification (Frost et al., 2006; 

Cairns and Yan, 2009), suggesting that recovery 

may ultimately be closer to Point 4 than Point 1. 

Figure 3-1. Conceptual figure showing hypothetical acidification and recovery patterns in a waterbody that has been 

acidified by atmospheric deposition (prepared by USGS).

Acid Impacted

Relative Measure of Pre-Acidication

Biological Community

Point 2

Point 3

Point 4

More Acidic Less Acidic

Healthy

Acid Impacted

Acidication

Recovery

to Date

Likely Range of

Future Recovery

Note: The two dashed lines 

reect the range of possible 

recovery trajectories for 

the biological community 

once 100% chemical 

recovery has been reached, 

as determined by ANC or 

another measure of stream 

acidity. The changes may or 

may not be linear with time 

as shown.

Point 1

Despite the concerns raised about limitations on 

recovery from acidification due to competition, 

dispersal, and other ecological factors that 

must be considered in addition to chemical 

thresholds, several studies in Europe and 

Canada have demonstrated biological recovery 

in aquatic communities that have become less 

acidic in response to decreases in acid deposition 

(Tipping et al., 2002; Findlay, 2003; Halvorsen 

et al., 2003). Additionally, studies in which 

lime has been added to waterbodies or soils to 

hasten recovery from acidification have shown 

trends towards a pre-acidification aquatic 

community (Raddum and Fjellheim, 2003), 

but the success of these efforts has sometimes 

required active restoration of some species 

(Driscoll et al., 1996; Clair and Hindar, 2005). 

There have been few studies in the United States 

on the recovery of aquatic biota over a period 

of gradually improving acidification status; 

this is largely because none of the existing U.S. 
Chapter 3 | Effects of Acid Deposition on Ecosystems

52

monitoring networks regularly provide such 

evaluations and because most available data 

originate from ad-hoc studies. Nonetheless, a 

few studies have evaluated aquatic communities 

by sampling waters over a period when acidity 

was declining, and the general conclusion is 

that there is little evidence to date of widespread 

recovery in these communities (Simonin et al., 

2005; Burns et al., 2008a; Warren et al., 2008).

Terrestrial Recovery

Few studies have tracked the status and potential 

recovery of terrestrial ecosystems during the period 

of decreasing acid deposition since the 1980s. This 

is due in part to the historical focus on aquatic 

acidification and a lack of understanding of the 

terrestrial impacts of acidification at the time 

that the Clean Air Act was amended in 1990. 

Since the early 1990s, scientific understanding of 

terrestrial ecosystem effects of acid deposition has 

increased greatly, and studies indicate continued 

degradation of soil base status (calcium and 

magnesium), as well as continued deleterious 

effects on sensitive species such as sugar maple and 

red spruce in the most acid-sensitive regions. 

As described previously, several studies have shown 

the link between acid deposition and losses of 

base cations from soil. Many of these studies have 

included data from re-sampled soil locations, 

generally documenting a lack of improvement or 

continued degradation of soil base cation status in 

the Adirondack Mountains and Allegheny Plateau 

of Pennsylvania (Bailey et al., 2005; Sullivan et al., 

2006b; Johnson et al., 2008; Warby et al., 2009). 

Furthermore, a clear link between the base cation 

status of soils and the health of sugar maple has 

been established in recent years (Bailey et al., 2004; 

Hallet et al., 2006; Long et al., 2009). The role 

of acidification in red spruce freezing injury was 

established by the early 1990s (McLaughlin et al., 

1993). Recent data have largely confirmed the 

findings of earlier studies on red spruce (Hawley et 

al., 2006) and have also found potential effects of 

acid deposition on other species of forest vegetation 

(Jenkins et al., 2007; Zaccherio and Finzi, 2007). 

Currently, no published studies can document an 

improvement in the status of terrestrial ecosystems 

in the eastern United States relative to acidification 

effects due to decreased emissions and deposition. 

This conclusion likely results from three factors: 

(1) few studies have evaluated terrestrial ecosystem 

health relative to acidification effects over time, 

(2) soils in the most acid-sensitive regions continue 

to acidify, and (3) the potential for lags in response 

in long-lived species such as forest trees.

Currently, no published studies can document 

an improvement in the status of terrestrial 

ecosystems in the eastern United States 

relative	to	acidification	effects	due	to	

decreased emissions and deposition.

Central Appalachian basic boulderfield forest, 

Shenandoah National Park (photo courtesy of Gary P. 

Fleming, Virginia Department of Conservation and 

Recreation). 

Current Effects of Atmospheric 

Deposition on Aquatic and Terrestrial 

Ecosystems in the Rocky Mountain 

Region of the United States

As shown in Chapter 2, the Rocky Mountain 

region of Colorado and adjoining states receives 

atmospheric deposition levels of sulfur and nitrogen 

that are generally less than half the deposition 

levels measured in mountainous parts of the 

eastern United States. Yet, the Colorado Rockies 

landscape is generally more sensitive to nutrient 

over-enrichment from atmospheric nitrogen 

deposition than landscapes in the eastern United 
NAPAP Report to Congress: An Integrated Assessment

53

States. This sensitivity is due to the steep slopes, th

or non-existent soil cover, short growing season, 

and less vegetation per unit landscape area (Clow 

and Sueker, 2000). Several studies have measured 

ANC values near or below zero during snowmelt 

in headwater streams, small headwater lakes, and 

ephemeral pools, and downward trends in ANC 

values were evident in some waters through the 

1990s (Caine, 1995; Williams and Tonnessen, 

2000; Campbell et al., 2004). However, the extent 

of waters that become episodically acidified in 

the Colorado Rockies is generally less than in the 

eastern United States, and no deleterious effects 

of acidification on fish have been documented.

in 

The character of nitrogen deposition is 

different in the Colorado Rockies than in 

the East, with ammonia nitrogen deposition 

generally exceeding that of nitrate nitrogen.

Loch Vale in Rocky Mountain National Park 

receives among the highest loads of atmospheric 

nitrogen deposition in the Front Range of the 

Colorado Rockies (photo courtesy USGS).

Atmospheric deposition of nitrogen has been 

of interest in the Rocky Mountains because 

both aquatic and terrestrial ecosystems in this 

region were historically nitrogen limited and 

have shown species shifts and growth responses 

at nitrogen deposition levels as low as 1.5 kg/

ha/yr (Wolfe et al., 2003; Baron, 2006). Several 

factors distinguish nitrogen deposition patterns 

in the Colorado Rockies from those in the eastern 

United States. First, the character of nitrogen 

deposition is different in the Colorado Rockies than 

in the eastern United States. Ammonia nitrogen 

(i.e., NH3-N) deposition, which generates greater 

acidity than does nitrate, generally exceeds that of 

nitrate nitrogen (i.e., NO3-N), especially at sites 

east of the Continental Divide in the Front Range 

of the Rocky Mountains (Heuer et al., 2000; Burns, 

2003; Beem et al., 2010). This pattern is believed to 

result from the deposition of emissions carried by 

upslope winds to the Front Range from agricultural 

sources that are located to the east of the mountains 

(Baron et al., 2004). As a result, atmospheric 

nitrogen deposition at many monitoring sites 

in the Colorado Rockies is not decreasing with 

decreases in power plant emissions in the western 

United States. Instead nitrogen loads are still 

increasing in some western locations. This pattern is 

especially driven by a tendency towards increasing 

trends in ammonia deposition (Burns, 2003).

Many effects on terrestrial and aquatic ecosystems 

from current levels of atmospheric nitrogen 

deposition have been demonstrated in the Rocky 

Mountain region. Over the past two to three 

decades, increases in nitrate concentrations and 

inorganic nitrogen loads in high-elevation streams 

and lakes have been documented in a few areas of 

the Colorado Rockies (Williams and Tonnessen, 

2000; Clow et al., 2003). However, throughout 

a wider range of high-elevation waters in the 

western United States, trends have been slight 

enough that other factors such as recent rainfall 

(Clow et al., 2003) and the melting of glaciers 

(Baron et al., 2009) can greatly affect nitrate 

trends. Studies in which historical atmospheric 

deposition levels were indirectly evaluated through 

measurements of nitrogen levels in lake sediment 

cores have concluded that nitrogen deposition 

over the Colorado Rockies and many western 

mountain ranges began to increase greatly after 

the 1950s. This trend is marked by shifts in lake 

plankton communities from low nutrient species 

to those indicative of an increased nutrient supply 

(Wolfe et al., 2003; Saros et al., 2005). The 

general conclusion across high-elevation waters of 
Chapter 3 | Effects of Acid Deposition on Ecosystems

54

the western United States is that as atmospheric 

nitrogen deposition increases from low levels of 

< 2 kg/ha/yr to higher levels, plankton biodiversity 

tends to decrease, and growth tends to increase as 

the aquatic-limiting nutrient shifts from nitrogen 

to phosphorus (Interlandi and Kilham, 2001; 

Bergstrom and Jansson, 2006). These shifts and 

changes in plankton communities are assumed to 

occur at a given location as atmospheric nitrogen 

deposition increases over time and spatially from 

regions of low atmospheric nitrogen deposition to 

regions of high atmospheric nitrogen deposition.

The changes described above for aquatic ecosystems 

are also paralleled by observed changes in terrestrial 

plant communities, especially those in highelevation alpine climatic zones of the Colorado 

Rockies. Increases in atmospheric nitrogen 

deposition tend to increase growth and decrease 

species diversity in alpine plant communities, with 

sedge and grass communities generally favored over 

shrubs (Bowman et al., 1995; Seastedt and Vaccaro, 

2001). Changes in vegetation nitrogen content 

of individual species can occur at atmospheric 

nitrogen deposition levels as low as 4 kg/ha/yr, 

and whole community changes were observed at 

values of 10 kg/ha/yr, as determined by nitrogen 

addition experiments (Bowman et al., 2006). 

Together, the results of studies highlight the high 

sensitivity of aquatic and terrestrial ecosystems 

in the Rocky Mountain region to atmospheric 

nitrogen deposition. Deleterious effects are generally

observed at deposition levels that are significantly 

less than those commonly measured in highelevation ecosystems of the eastern United States. 

These results are pertinent to the management 

of ecosystems in this region because of the large 

number of National Parks (e.g., Rocky Mountain, 

Grand Teton, Yellowstone) that are managed as 

Class 1 Wilderness areas under Clean Air Act 

provisions to prevent significant deterioration of 

ecosystems (Porter and Johnson, 2007). Thus, a 

more active Federal role in land management and 

ecosystem protection from air pollutant deposition 

may be necessary in the large tracts of Class 1 

wilderness land that are present in the west.

Much of the Grand Teton National Park is managed as a 

Class 1 wilderness area (photo courtesy of NPS).

Current Effects of Atmospheric 

Nitrogen Deposition on Terrestrial 

and Aquatic Ecosystems in the 

Far Western United States

In many areas of the far western United States (lands 

to the west of the Rocky Mountains, especially 

the West Coast states), nitrogen loads to terrestrial 

systems have historically been low compared 

with loads to terrestrial systems in the eastern 

United States. The native plant communities (e.g., 

coastal sage scrub, desert scrub, annual grasslands) 

are adapted to thrive in low-nitrogen environments. 

When nitrogen loads to these habitats increase, the 

native plant and soil communities have difficulty 

competing with invasive species that are better 

adapted to more nitrogen-rich environments. In 

California, nitrogen deposition levels as low as 

5–8 kg/ha/yr have major effects on epiphytic lichens 

(i.e., lichens that grow on trees and shrubs) and 

contribute to changes in native plant communities 

by enhancing invasion by exotic annual grasses 

(Fenn et al., 2003a; Weiss, 2006). In coastal sage 

scrub, desert scrub, and annual grasslands in 

California, nitrogen deposition favors non-native 

grass invasions, which can lead to altered plant 

communities. In southern California chaparral 

(Egerton-Warburton et al., 2001) and coastal sage 

scrub ecosystems (Egerton-Warburton and Allen, 

2000), the diversity of the mycorrhizal fungi 

communities that grow in association with roots 

also is negatively impacted by nitrogen deposition. 

Shifts in these plant communities are of additional 

concern because the invasive vegetation that results 
NAPAP Report to Congress: An Integrated Assessment

55

from increased nitrogen loads also may increase the 

risk of wildfires due to increased fuel availability. In 

the desert scrub ecosystems of Joshua Tree National 

Park, invasive grasses can provide sufficient fuel 

load to carry fire, particularly in a wet year. This 

increased fuel load puts the Park’s native vegetation 

at risk, including its namesake, the Joshua tree. 

Fire occurrence in the desert is a rare event in the 

absence of the invasive grasses (Allen et al., 2009). 

Desert scrub ecosystem in Joshua Tree National Park 

may be impacted by nitrogen deposition (photo 

courtesy of USGS).

Increased nitrogen deposition also can impact 

forests in the far western United States. Nitrogen 

deposition effects, in combination with the 

physiological effects of ozone, decrease the amount 

of roots and increase susceptibility of ponderosa 

pine to bark beetle attack and mortality (Jones et 

al., 2004). Forest sustainability is thus threatened, 

and fire risk is increased (Grulke et al., 2009). 

Studies have been conducted to assess various 

methods of reducing nitrogen saturation symptoms

Using fire as a management technique to reduce 

excess nitrogen was found to be ineffective in 

chaparral ecosystems, presumably because of the 

limited capacity of fire to reduce the large nitrogen 

stores in the mineral soil of these ecosystems 

(Meixner et al., 2006). Similarly, in forests affected 

by excess nitrogen, a combination of periodic 

nitrogen release by fire and significant reductions 

in nitrogen deposition are needed to mitigate the 

effects of nitrogen accumulated from long-term 

nitrogen deposition (Gimeno et al., 2009).

. 

Nitrogen deposition effects on the lichen 

community are pronounced and widespread 

in the forests, oak woodlands, and chaparral of 

California, as evidenced by major shifts from 

dominance of lichen species that thrive in lowacid, low-nitrogen environments to communities 

dominated by lichen species that thrive in acidic, 

high-nitrogen environments (Fenn et al., 2008; 

Jovan, 2008; Jovan and McCune, 2005). Similar 

effects of nitrogen in lichen communities are well 

documented in forests near urban and agricultural 

emissions source regions in Washington and 

Oregon and in the Columbia River Gorge corridor 

(Fenn et al., 2007; Geiser and Neitlich, 2007). 

In the streams and lakes of the far western United 

States, increased nitrogen deposition can alter 

lake and stream chemistry. At moderate nitrogen 

deposition levels of 10–14 kg/ha/yr for chaparral 

catchments and of 17 kg/ha/yr for forested 

catchments in the southwestern Sierra Nevada 

and in the Transverse Ranges near Los Angeles, 

high concentrations of nitrate are found in stream 

water. In chaparral, forested, and high-elevation 

wilderness catchments in the Los Angeles air basin, 

up to 40% of the nitrate exported in stream water 

was directly deposited from the atmosphere and 

transported through the system without biological 

assimilation (Michalski et al., 2004), clearly 

indicating that the source of this excess nitrogen 

was air pollutant deposition in these watersheds.

High-elevation watersheds in the Sierra Nevada 

appear to export nitrate during early snowmelt, 

even with relatively low atmospheric nitrogen 

deposition similar to the high tendency to export 

nitrate discussed above for the Colorado Rockies. 

Although the role of nitrogen deposition on nitrate 

leaching in the high-elevation Sierra Nevada is not 

entirely clear, nitrogen deposition does not appear 

to have as severe an impact on lake and stream water

chemistry, particularly in Sierra Nevada catchments, 

where soils tend to be thicker than those of the 

Colorado Front Range. Differing responses 

between the two regions may be due to higher 
Chapter 3 | Effects of Acid Deposition on Ecosystems

56

nitrogen deposition in the Colorado Rockies or to 

environmental differences. For example, the greater 

extent of soil freezing in the Colorado Rockies 

may reduce the nitrogen retention capacity of 

these watersheds (Sickman et al., 2002). However, 

watersheds in the Sierra Nevada are expected to 

be sensitive to increases in nitrogen deposition in 

the seasonal snowpack. This expectation is based 

on the large pulses of nitrate that are flushed to 

surface waters during snowmelt-driven spring 

runoff (Sickman et al., 2003); a short growing 

season; the limited soil and vegetation cover; and 

the limited biological nitrogen demand during the 

early spring runoff period (Fenn et al., 2003b). 

Effects of Atmospheric Nitrogen 

Deposition on Coastal Estuaries

Eutrophication and associated hypoxia is 

widespread in estuaries of the United States (Figure 

3-2),including Long Island Sound, the Chesapeake 

Bay, and the Gulf of Mexico near the Mississippi 

River Delta (Diaz, 2001; Kemp et al., 2005). 

Nitrogen and phosphorus are the nutrients that 

commonly limit algal growth in coastal ecosystems. 

The role of nitrogen in estuarine eutrophication 

and hypoxia is often attributed to a combination 

of the delivery of excess nitrogen by rivers and the 

direct atmospheric deposition of nitrogen onto 

the water surface. These same mechanisms also 

deliver nitrogen to the open ocean, and atmospheric 

nitrogen deposition has been shown to affect the 

growth of marine plankton (Krishnamurthy et al., 

2010). The potential sources of nitrogen to estuaries 

include runoff from fertilizer and animal waste 

applied to agricultural landscapes, human waste 

in urban/suburban landscapes, and atmospheric 

nitrogen deposition. The relative role of atmospheric 

nitrogen deposition to total nitrogen loads to coastal 

estuaries differs widely, from about 15% to 25% in 

the Mississippi Delta region of the Gulf of Mexico, 

to 20% to 50% in Chesapeake Bay and Tampa Bay, 

and as much as 60% to 80% in some coastal river 

estuaries of northern New England (McIsaac et al., 

2001; Boyer et al., 2002; Poor et al., 2006; Whitall 

et al., 2007; Alexander et al., 2008; Howarth, 

2008). The relative contributions of nitrate and 

ammonia as sources of atmospheric nitrogen to 

estuaries also vary widely across coastal regions of 

the United States. Nitrate is generally the dominant 

form of nitrogen in atmospheric deposition across 

the United States; however, ammonia can be the 

dominant atmospheric species of nitrogen in parts 

of estuarine watersheds, particularly those with 

abundant agricultural land (Lawrence et al., 2000; 

Whitall et al., 2003). The most recently available 

data indicate that the majority of U.S. estuaries 

show signs of eutrophication (Bricker et al., 2008), 

and the number of impacted estuaries in the United 

States and globally has increased greatly since the 

1960s (Diaz and Rosenberg, 2008). Although 

most U.S. coastal waters have not shown changes 

in nutrient loads or trophic status since the 1990s 

(Alexander and Smith, 2006), about 20% to 25% 

of these coastal waters show improved trophic 

status in recent decades (Alexander and Smith, 

2006; Bricker et al., 2008), most notably Tampa 

Bay. The role of decreased loads of atmospheric 

nitrogen deposition to observed decreases in 

estuarine eutrophication is likely only minor, but is 

not well quantified (Greening and Janicki, 2006).

Figure 3-2. Map of major estuaries in the United 

States (prepared by U.S. EPA).

The contribution of headwater streams to the 

total nitrogen loads in rivers can be substantial. 

For example, small streams account for more than 

two-thirds of the water volume and approximately 

two-thirds of the nitrogen delivered to large rivers 

in the Northeast (Alexander et al., 2007). Despite 

a higher removal rate (denitrification) of nitrate 

from headwater reaches than from large river 

channels during transport (Alexander et al., 2008), 

the colder temperatures at the higher elevations 

of the headwaters act to slow this rate of loss 

(Schaefer and Alber, 2007). Therefore, a substantial 
NAPAP Report to Congress: An Integrated Assessment

57

amount of the nitrogen deposited on upland 

forested areas from the atmosphere may later be 

exported to larger rivers and estuaries. These studies 

emphasize the connection of upland ecosystems to 

coastal ecosystems and indicate that the effects of 

atmospheric nitrogen deposition extend beyond 

acidification in sensitive upland landscapes.

The issue of ocean acidification has emerged in 

recent years, with the observation that rising levels 

of carbon dioxide (CO2) in the atmosphere are 

causing decreases in ocean pH and alkalinity, with 

numerous implications for the health of marine 

ecosystems (Caldeira and Wickett, 2003). Some 

recent studies have explored the role of atmospheric

nitrogen and sulfur deposition on ocean 

acidification and concluded that acid deposition 

contributed no more than a few percent to ocean 

acidification trends in recent years (Bates and Peters,

2007). However, the contribution of acid depositio

to marine pH declines is likely greater in coastal 

regions closer to continents (Doney et al., 2007). 

Further scrutiny of the role of acid deposition on 

the pH and alkalinity of coastal ecosystems may be 

needed in the future given the serious consequences 

of ocean acidification that are currently forecast.

n 

Eutrophication and Hypoxia

Eutrophic waters are those with an 

overabundance of nutrients, such as nitrogen 

or phosphorus. As nutrient levels in the water 

increase, the growth of algae and plankton 

increases. When these organisms die, they 

are decomposed by bacteria that use the 

dissolved	oxygen	in	the	water.	“Dead	zones”	

characterized by hypoxia (i.e., low levels of 

dissolved oxygen in the water) or anoxia (i.e., 

no dissolved oxygen in the water) may result, 

harming	or	killing	fish,	crabs,	shrimp,	and	

other	organisms.	Benthic	(bottom-dwelling)	

organisms such as oysters that are unable 

to move to more oxygenated waters are 

particularly susceptible.

Emerging Issues at the 

Interface of Science and Policy 

in Managing the Effects of Acid 

Deposition on Ecosystems 

Critical Loads

Critical loads, which were introduced in Chapter 2, 

can be viewed as providing a synthesis of scientific 

knowledge about thresholds of harm to or alteration 

of ecosystems that can inform the development, 

implementation, and assessment of air quality 

policy and the management of sensitive ecosystems. 

Critical loads have been widely used for large-scale 

applications and policy purposes in Europe and 

Canada since the 1980s (Jeffries and Ouimet, 2005; 

Spranger et al., 2008), but historically have not 

been widely used for these purposes in the United 

States (Porter and Johnson, 2007). For example, 

critical load maps for soils and waters are produced 

by several European countries that are parties to 

the Convention on Long-Range Transboundary Air 

Pollution. Although a signatory to this Convention, 

the United States did not commit to the provisions 

regarding critical loads contained in the Convention 

(see recent discussion in Spranger et al., 2008). 

Academic critical load studies in the United States 

also date to the 1980s (Henriksen and Brakke, 

1988), and many studies continue to be published 

(Baron, 2006; Fenn et al., 2008; Sullivan et al., 

2008); however, most past work has been of 

limited geographic scope. Recently published 

studies provide examples of larger-scale assessments 

of critical loads in the United States, including 

nationwide critical load maps for forest soils 

(McNulty et al., 2007) and a review of published 

studies throughout the United States for empirical 

critical loads of nitrogen (Pardo et al., 2011).

Critical loads can be calculated for any species of 

concern or for a sensitive biological community 

within an aquatic or terrestrial ecosystem if a 

threshold of atmospheric deposition can be 

identified below which negative biological effects 

on the species or alteration of the community are 

not expected to occur. Because various species in 

an ecosystem may have differing sensitivities to air 

pollutant loads, a large number of species-specific 

or community-specific critical loads are likely to 

exist for a given location (Sullivan et al., 2008). 
Chapter 3 | Effects of Acid Deposition on Ecosystems

58

Critical loads are often linked to thresholds at 

which the organism experiences a certain biological 

effect, which may range from stunted growth to 

reduced reproductive levels to increased mortality. 

Critical loads are usually calculated in reference to 

a chemical quantity believed to provide an index of 

acidification or nutrient over-enrichment. Potential 

indices include ANC levels or nitrate concentrations 

in a waterbody, or the base cation saturation level of 

a soil. Multiple critical loads and ecological effects 

end points provide policy makers or land managers 

with multiple options to determine the level of 

protection that is warranted based on the species 

or communities of greatest concern at a given 

location. This is part of the process of establishing 

a target load, which may differ from the strictly 

science-based critical load. The target load reflects 

inclusion of economic, social, cultural, or other 

considerations to establish the protection level of 

the ecosystem in question, as well as the load that 

is deemed “achievable” by managers or regulators 

(Porter et al., 2005; Porter and Johnson, 2007).

In addition to a variety of critical loads applicable 

for each ecosystem based on the species and 

biological effects being considered, there are 

different types of critical loads and different 

methods for establishing critical loads. Critical 

loads can be determined by an empirical approach 

based on field- or laboratory-based studies that 

define a chemical threshold of sensitivity for a 

species or community. A recent example is the 

establishment of a critical load of 1.5 kg/ha/yr 

of atmospheric nitrogen deposition to protect 

against eutrophication (nutrient over-enrichment) 

in lakes in the Rocky Mountain National Park. 

This load was established based on evidence 

from lake sediment cores that showed a shift in 

phytoplankton species as nitrogen deposition 

values rose above 1.5 kg/ha/yr in the 1960s 

(Porter and Johnson, 2007). Critical loads may 

also be determined by a mathematical expression, 

such as the Steady-State Mass Balance model for 

soils. This model compares the concentrations of 

neutralizing base cations (e.g., calcium, magnesium) 

to those of potentially toxic acidifying cations 

(e.g., aluminum) in soils (Sverdrup and deVries, 

1994). Both the empirical and mathematical 

expression approaches develop steady-state critical 

loads based on an assessment of current levels of 

atmospheric deposition relative to some threshold 

of harm or change. Critical loads can also be 

calculated using a dynamic approach that involves 

more sophisticated modeling of time-varying 

patterns of emissions, deposition, soil chemistry, 

and lake or stream chemistry. This approach 

often uses some of the acidification models (e.g., 

MAGIC or photosynthetic evapotranspirationbiogeochemical [PnET-BGC]) described in the 

text box and can provide information on the time 

expected to reach a threshold based on a given 

emissions policy as it is implemented over time 

(Sullivan et al., 2008). The dynamic critical load 

approach is especially valuable for evaluating 

the success of air quality policies and programs 

as they are implemented over multiple years.

Despite the lack of a long history of applying 

critical loads to inform air pollutant policies in 

the United States, interest in the use of critical 

loads in the United States has increased in recent 

years with the advent of a Critical Load Science 

Committee within NADP (NADP, 2009), several 

recent workshops and meetings on this topic, 

and several publications exploring greater use of 

critical loads as a policy-relevant environmental 

assessment tool (Porter et al., 2005; Fisher et al., 

2007; Burns et al., 2008b; Glavich and Geiser, 

2008). Recently published studies have generally 

found that for many regions well known for their 

sensitivity to either acidification or excess nitrogen 

deposition (e.g., the Adirondack Mountains of 

New York, the central Appalachian region of 

Virginia, the Front Range of Colorado), the current 

levels of atmospheric deposition generally exceed 

the critical loads for many species of concern 

(Baron, 2006; McNulty et al., 2007; Sullivan et 

al., 2008). Case studies exploring the application 

of the critical load approach for the Adirondack 

Mountains and the central Appalachian region 

are also included in Chapter 2 and Chapter 4 

of this report. Finally, consideration of two 

key factors is warranted when using scientific 

information on critical loads to inform public policy 

implementation: (1) critical loads can have high 

uncertainty that stems from model assumptions 

and data, as well as uncertainty in the threshold 

itself (Skeffington, 2006; Li and McNulty, 2007), 
NAPAP Report to Congress: An Integrated Assessment

59

and (2) other variables such as forest harvesting 

and climate change will affect critical loads in 

addition to atmospheric deposition (Watmough 

et al., 2003). When critical load calculations 

are being used to inform air pollutant policies, 

scientific uncertainty must be properly considered, 

along with the role of other mitigating variables.

Atmospheric Deposition 

and Ecosystem Services

In addition to the health benefits of the ARP 

discussed in Chapter 1 of this report, benefits to 

ecosystem services also might be expected, and 

these benefits can be evaluated by considering the 

varied material and intangible goods and services 

by which ecosystems support human well-being 

(PCAST, 2011). Ecosystem services can be 

defined as the sum total of benefits that humans 

receive from the resources and processes of natural 

ecosystems (Corvalan et al., 2005). These benefits 

range from extractive services or provisioning (e.g., 

wood products, minerals, water) to regulating 

services (e.g., climate regulation, nutrient cycling) 

to cultural services (e.g., outdoor recreation). The 

ability of ecosystems to provide these services over 

the long term is impacted by disturbances (Beier 

et al., 2008), including atmospheric deposition. 

The response of ecosystems to disturbances, such 

as acid deposition, depends on the resiliency of the 

ecosystem in question. Over the past two decades, 

researchers have begun to establish the value of 

these various ecosystem services (Costanza et al., 

1997) and to develop approaches for valuating 

both the direct use-related benefits, as well as 

the non-use benefits of ecosystems; most studies 

indicate these non-use benefits exceed those of the 

more traditional market-based benefits derived 

from resource extraction (Walsh et al., 1990).

Models Used to Investigate the Ecosystem Effects of Acid Deposition

Models play a pivotal role in understanding the 

ecosystem effects of acid deposition, as well as 

helping to inform air quality policy. Below are the 

principal models that have been used to study 

ecosystem effects.

Initial Models—These models were simple and 

used an analysis of surface water chemistry based 

on charge balance or titration by strong acids 

(Henriksen, 1979).

Dynamic Models—Later, more formal 

computer models were developed that simulate 

biogeochemical processes and hydrology 

occurring	in	catchments	(see	the	figure	on	the	

following page). The dynamic models listed all 

share an ability to make predictions about future 

concentrations and loads of chemical constituents 

(e.g., pH, ANC) in lakes and streams given the 

chemistry of atmospheric deposition. 

•	 Integrated	Lake-Watershed	Acidification	Study	

(ILWAS)	model	(Gherini	et	al.,	1985)

•	 Birkenes	model	(Christophersen	et	al.,	1982)

•	 Model	of	Acidification	of	Groundwater	in	

Catchments	(MAGIC)	(Cosby	et	al.,	1985c)

•	 Photosynthetic	evapotranspirationbiogeochemical	(PnET-BGC)	(Gbondo-Tugbawa	

et al., 2001).

All of the dynamic models can be combined with 

atmospheric models, such as the Community 

Multiscale Air Quality model (CMAQ), that 

estimate atmospheric deposition to landscapes 

from present emission levels or future emission 

scenarios (Schwede et al., 2009). 

(continued)
Chapter 3 | Effects of Acid Deposition on Ecosystems

60

Conceptual diagram of the biogeochemical processes and chemical stores of 

the PnET-BGC model (figure courtesy of Charles Driscoll, Syracuse University).

Models Used to Investigate the Ecosystem Effects of Acid Deposition (continued)

Wet

Deposition

Dry

Deposition Climatic data

• Solar radiation

• Precipitation

• Temperature

PnET

 Water balance

 Photosynthesis

 Living biomass

 Litterfall

Net Mineralization

Uptake

BGC

• Aqueous reactions

• Surface reactions

 • Cation exchange

 • Adsorption

 • Humic binding

 • Aluminum dissolution/precipitation

Shallow water ow

Weathering Deep water ow

BGC – Surface water

Aqueous reactions

Other dynamic models, such as the following, have 

been applied more commonly in ecosystems of 

the western United States: 

•	 DayCent/PHREEQC	(DayCent-Chem)	

(Hartman et al., 2007)

•	 Alpine	Hydrochemical	Model	(AHM)	(Meixner	

et al., 2004).

Steady State Models—These are another class 

of	models	that	focus	only	on	a	steady-state	

approximation of either water or soil chemistry 

(Henriksen	and	Posch,	2001).	Steady-state	models,	

such as those listed below, are computationally 

simpler and demand less input data to run:

•	 Steady-State	Water	Chemistry	(SSWC)	

•	 Simple	Mass	Balance	Equation	(SMBE).	

Several recent surface water chemistry modeling 

studies have explored likely future ecosystem 

responses to existing air pollutant policies 

(e.g., Title IV),	along	with	some	future	emissions	

scenarios (Chen and Driscoll, 2004; Sullivan et al., 

2004;	Chen	and	Driscoll,	2005a;	Chen	and	Driscoll,	

2005b;	Lawler	et	al.,	2005;	Sullivan	et	al.,	2008).	

These	modeling	studies	not	only	generally	confirm	

the	success	of	Title	IV	at	improving	the	acid-base	

status of surface waters, but also highlight regional 

differences and point out that additional emission 

reductions will be needed to realize the goal of full 

ecosystem recovery to conditions resembling 

pre-acidification	conditions.	Chapter	4	of	this	

report includes a modeling analysis in which 

MAGIC	was	used	to	assess	aquatic	ecosystem	

recovery in several regions of the eastern United 

States as a result of implementing current policies 

and programs in comparison to results achieved as 

a result of implementing three alternative future 

emission-reduction	scenarios.
NAPAP Report to Congress: An Integrated Assessment

61

A recent review discusses and summarizes studies 

on estimating the monetary value of the effects 

of sulfur and nitrogen emissions and deposition 

on ecosystems (Krupnik and Siikimaki, 2008). 

The studies cited can be divided into those that 

estimate the monetary value of terrestrial, aquatic, 

and transitional (mainly wetland) ecosystems. 

Several studies have generally estimated the value 

of adverse ecosystem effects, but many fewer 

studies have attempted to relate specific adverse 

effects that can be directly attributed to sulfur 

and nitrogen deposition. Some studies have 

estimated the value of benefits from minimizing 

or eliminating ecosystem disturbances such 

as insect defoliation of forests (Kramer et al., 

2003) and estuarine eutrophication (Morgan 

and Owens, 2001), in which acid deposition 

is clearly one contributing factor among many. 

However, the monetary role of acid deposition 

cannot be estimated in most of these studies.

The greatest challenge in developing specific data on 

the economic benefits of changes in acid deposition 

due to emission reductions lies with the availability 

of comprehensive scientific evidence that defines 

the extent and magnitude of the adverse effects that 

can be directly attributed to acid deposition from 

among multiple ecosystem stressors (Krupnik and 

Siikimaki, 2008). In one notable study, Banzhaf et 

al. (2006) estimated ecological benefits of a policy 

that would substantially decrease adverse effects 

of acid deposition in the Adirondack Park of New 

York. These benefits are in the range of about $300 

million to $1.1 billion annually for the Adirondack 

Park only and do not consider ecological benefits 

in other regions of the United States from these 

deposition reductions. As described in Evans 

et al. (2008), researchers are applying economic 

methods to show the total value people place on 

ecological improvements to important resources, 

such as the Adirondack Park, expected from 

further reductions in acid deposition. However, as 

discussed in Chapter 1 of this report, no studies 

have comprehensively evaluated the monetary 

benefits to ecosystem services of the ARP across all 

ecosystem types that would be sufficient to make 

comparisons to overall health benefits and costs of 

the ARP. Significant future analytical work and basic 

ecological and economic research is needed to build 

a sufficient base of knowledge and data to support 

an adequate assessment of ecological benefits.

Forest harvesting is an example of an ecosystem service 

that provides monetary benefit to land owners (photo 

courtesy of Douglas Burns, USGS).

What are the Implications of 

Ongoing and Future Changes 

in Climate and the Carbon 

Cycle for the Effects of Acid 

Deposition on Ecosystems?

The Earth has warmed by an average of 0.74° C 

between 1906 and 2005, and there is a high 

likelihood that recent warming over the past several 

decades has been driven in large part by increases 

in atmospheric concentrations of CO2 and other 

greenhouse gases (IPCC, 2007). Under a range of 

likely future scenarios of greenhouse gas emissions 

and a range of climate model predictions, the 

Earth is likely to warm by another 1.4°–5.8° C 

during the twenty-first century (IPCC, 2007). 

Ongoing and future predicted changes in the 

climate and the carbon cycle have numerous 

implications for the response of ecosystems to 

atmospheric deposition of sulfur and nitrogen. In 

addition to warming air temperatures, a wide array 

of other environmental variables are predicted to 

change with the climate, including water availability 

and the rates of many nutrient cycling processes. 

These additional environmental changes and their 

effects are likely to vary by region (Boisvenue and 

Running, 2006). In short, ongoing and future 

predicted climate change is likely to greatly affect 
Chapter 3 | Effects of Acid Deposition on Ecosystems

62

the interactions of sulfur and nitrogen deposition 

with ecosystems in a complex manner that does 

not easily lend itself to simple summaries and has 

high uncertainty (McLaughlin and Percy, 1999). 

For example, although the effects of an average 

increase in air temperature are likely to induce 

widespread changes in ecosystems, an increase in 

the magnitude and frequency of a wide variety of 

climatic events (e.g., floods, droughts, fires) also is 

expected. Some have suggested that these events 

are likely to affect ecosystems to a greater extent 

than long-term average changes in temperature 

(Dale et al., 2001; Jentsch et al., 2007). Despite 

these complexities, a brief assessment of likely 

interactions between ecosystems, acid deposition, 

and climate change is warranted in this current 

assessment given the potential magnitude and 

multi-faceted impacts of the projected changes. 

This section focuses on ecosystem effects and 

does not discuss the interactions between climate 

change and greenhouse gases or how these 

interactions may affect the transport, atmospheric 

interactions, and atmospheric deposition of sulfur 

and nitrogen. These atmospheric changes have 

been discussed elsewhere (Tagaris et al., 2008) 

and are beyond the scope of this assessment.

Ongoing and future predicted climate change 

is likely to greatly affect the interactions 

of sulfur and nitrogen deposition with 

ecosystems in a complex manner that does 

not easily lend itself to simple summaries and 

has high uncertainty.

Implications of Climatically 

Driven Changes in the Species 

Composition of Ecosystems

Perhaps the greatest impact of climate change is 

likely to be wholesale changes predicted for the 

biological communities found in many ecosystems 

sensitive to acid deposition effects, particularly 

those at high elevations (Ibanez et al., 2007). For 

example, a migration northward and to higher 

elevations is predicted for many tree species 

under most current climate change forecasts as 

they are replaced by species better adapted to a 

warmer climate (McKenney et al., 2007; Lenoir 

et al., 2008). Evidence indicates that migration 

of lower elevation forests to higher elevations is 

already occurring in Vermont, suggesting this 

shift is ongoing in mountainous areas that have 

warmed in recent decades (Beckage et al., 2008). 

Similar cold-to-warm water species shifts are also 

predicted for aquatic ecosystems (Mohseni et al., 

2003; Durance and Ormerod, 2007). The rate at 

which these ecosystem changes will occur is not 

well known and is likely to reflect interactions with 

other contemporaneous changes, such as those in 

water availability, light, insect defoliation, land 

use, and air pollutant deposition. Overall, the 

evidence suggests that many species may reach an 

out-of-equilibrium state with respect to a future 

warmer climate (Higgins and Harte, 2006; Mohan 

et al., 2009), and this state may result in increased 

sensitivity of many terrestrial and aquatic ecosystems 

to the effects of other disturbances, such as sulfur 

and nitrogen deposition. A European study (Alcamo 

et al., 2002) concluded that climate change is likely 

to have little effect on the sensitivity of forested 

ecosystems to air pollutant deposition. However, 

the weight of evidence from other studies indicates 

that, despite some uncertainty in the effects of 

climate change on ecosystems (Aber et al., 2001), 

ongoing and likely future climate change will 

provide an added stressor to many ecosystems in 

the United States. Furthermore, several studies 

indicate that the combination of atmospheric 

deposition of sulfur and nitrogen with high ozone 

concentrations will act as additional ecosystem 

stressors in many regions, such as the eastern United 

States (McLaughlin and Percy, 1999; Mohan et 

al., 2009). The increases in extreme events, such 

as droughts and fires, which are predicted with 

future climate change, will likely act as additional 

stressors to many of these ecosystems, with 

greater effects expected in the western United 

States (Dale et al., 2001; Knapp et al., 2008).

Climate Warming Will Affect 

Many Temperature-Sensitive 

Biogeochemical Processes in 

Ecosystems

Many biogeochemical processes affect the cycling 

of atmospherically deposited sulfur and nitrogen 

in ecosystems and play a pivotal role in the 
NAPAP Report to Congress: An Integrated Assessment

63

acidification of ecosystems, including mineral 

weathering, nitrification, and sulfur and nitrogen 

mineralization. The rates of nearly all of these 

processes are likely to be affected by changes 

in climate. For example, the rates of mineral 

weathering and nitrification increase with increasing 

temperature if all other factors are held constant 

(White et al., 1999; Dalias et al., 2002). Studies 

have shown widely varying biogeochemical 

responses to past and predicted climate warming 

in the United States, and these varying responses 

reflect interactions with precipitation, snow and 

ice, and seasonal variations. For example, Campbell 

et al. (2009) predict higher nitrate concentrations 

in soil water and streams at Hubbard Brook, NH, 

due to warming-induced increases in nitrogen 

mineralization and nitrification, yet they also 

predict decreased rates of mineral weathering 

due to decreases in soil moisture. At the Rocky 

Mountain National Park in the western United 

States, stream nitrate concentrations have been 

increasing in recent years due to the melting of 

glaciers and snow fields by increased warming in 

summer through fall and the leaching of nitrate 

that occurs when the underlying soils are exposed 

to microbial processes (Baron et al., 2009). These 

results highlight the sensitivity of long-term stream 

water quality to changes in snowfall, snowmelt, and 

the mass of glaciers in high-elevation watersheds of 

the western United States. Data from the European 

Alps show a recent trend towards increasing solute 

concentrations in high-elevation lakes, consistent 

with accelerated mineral weathering rates driven by 

warmer air temperatures and less-extensive snow 

cover that lasts for a shorter duration (Mosello 

et al., 2002; Rogora et al., 2003). Increased mineral 

weathering rates would likely result in a greater 

ability of soils to neutralize acid deposition and 

might accelerate recovery in sensitive ecosystems 

when accompanied by decreasing atmospheric 

deposition rates of sulfur and nitrogen. However, 

a recent study in the Colorado Rockies showed 

that regions underlain by sulfide minerals such 

as pyrite may experience greater acidification 

with warmer air temperatures due to increased 

mineral weathering rates and the subsequent 

formation of sulfuric acid (Mast et al., 2010).

For a range of likely twenty-first century climate 

change scenarios in Europe, Posch (2002) 

concluded that ecosystems will generally become 

less sensitive to acid deposition, primarily due 

to increasing rates of mineral weathering in soils 

driven by warmer temperatures. The amount of 

land in Europe subjected to sulfur and nitrogen 

deposition loads in excess of the critical load for 

acidification is likely to decrease with climate 

change, except in areas predicted to become more 

arid. Decreases in the exceedance of critical loads 

for nitrogen-driven nutrient over-enrichment 

are also predicted with future climate change in 

Europe (Posch, 2002), but these decreases were less 

than those modeled for decreases in critical load 

exceedances for acidification. In contrast, existing 

studies in the United States predict higher nitrate 

concentrations (which should lead to greater critical 

load exceedances for nutrient over-enrichment) 

with future climate change (Campbell et al., 2009; 

Wu and Driscoll, 2010). These opposing forecasts 

for expected future nitrate chemistry in surface 

waters indicate the current high uncertainty about 

the role of climate change and point to a need for 

additional research and improved models to explore 

this important issue (Spranger et al., 2008). 

Climate change studies in Loch Vale have documented 

increasing air temperatures, earlier snowmelt, and earlier 

runoff. These changes may have important impacts on 

water availability, water quality, and ecosystem function 

(photo courtesy of USGS).
Chapter 3 | Effects of Acid Deposition on Ecosystems

64

Implications of Changes in 

Carbon Dioxide Concentrations 

in the Atmosphere

One of the fundamental factors that affects the rates 

of photosynthesis and terrestrial ecosystem growth 

is the availability of CO2. Global atmospheric 

CO2 concentrations have increased by nearly 40% 

since the mid-eighteenth century and are expected 

to double during the twenty-first century (IPCC, 

2007). The ecosystem “CO2 fertilization effect” has 

been much discussed in the literature (Tans et al., 

1990; Nowak et al., 2004). Indeed, measurements 

and satellite observations of forested ecosystems, 

as well as CO2 manipulation experiments, have 

generally reported increasing net ecosystem growth 

(primarily in forests) as CO2 concentrations 

have increased in temperate regions of the world, 

although there have been fewer monitoring-based 

studies in the United States than in Europe (Norby 

et al., 2005; Boisvenue and Running, 2006). The 

ecological responses have not been consistent, 

however, in all studies and across all regions. A 

wide variety of factors (e.g., disturbance history, 

water availability, solar radiation), in addition to 

experimentally altered CO2 concentrations, can 

contribute to these diverse patterns in net terrestrial 

ecosystem growth trends (Nemani et al., 2003; 

Boisvenue and Running, 2006; Potter et al., 2006). 

Research at the Forest-Atmosphere Carbon 

Transfer and Storage (FACTS-I) facility in the Duke 

Forest has shown that nitrogen deposition can 

affect the uptake of atmospheric carbon dioxide as 

forests grow (photo courtesy of U.S. Department 

of Energy’s Brookhaven National Laboratory).

There are several challenges in trying to quantify 

the role of CO2 in forest growth. Many factors 

generally co-vary across sites, and experiments 

necessary to understand the interacting effects of 

these multiple factors have not been undertaken 

(Norby and Luo, 2004; Rustad, 2008). Also, the 

nearly impossible task of designing experiments 

that adequately consider variations in the numerous 

factors that influence ecosystem growth has been 

noted (Aber et al., 2001; Boisvenue and Running, 

2006). When variations in atmospheric sulfur 

and, especially, nitrogen deposition are added to 

this mix of factors, the difficulty in attributing 

cause and effect to observed changes in ecosystems 

increases further (Bytnerowicz et al., 2007), and 

some argue, can only be interpreted through the 

use of models (Aber et al., 2001). Nonetheless, 

results from a CO2 enrichment experiment that 

simulated likely twenty-first century concentrations 

showed increased soil acidification and increased 

rates of mineral weathering, suggesting that 

important interactions with terrestrial ecosystem 

effects of atmospheric deposition may occur in 

the future (Andrews and Schlesinger, 2001).

Changes in the Water Cycle are Key 

to Understanding Ecosystem Effects

Under various likely twenty-first century climatechange scenarios, most climate models predict an 

overall increase in global precipitation of about 

1–3% per degree of temperature increase (Held and 

Soden, 2006). Recent observations and satellite 

measurements generally support an even greater 

increase in precipitation of about 5–10% per 

degree of warming (Wentz et al., 2007; Zhang et 

al., 2007). Such a change is expected to increase 

the wet deposition of sulfur and nitrogen in the 

eastern United States if all other factors remain 

the same (Civerolo et al., 2008). However, the 

prediction of climate “wetting” is less certain than 

the prediction of climate warming, and changes in 

precipitation patterns are predicted to vary widely 

over the United States, with forecasts for a wetter 

climate in the eastern United States and for a drier 

climate in large parts of the western United States 

(Milly et al., 2005). These drier climate patterns are 

already evident in the western United States, which 

is exhibiting smaller winter snowpacks and earlier 

snowmelt in mountainous areas (Mote et al., 2005). 
NAPAP Report to Congress: An Integrated Assessment

65

Earlier snowmelt has also been noted in upland 

watersheds in the Northeast (Burns et al., 2007). 

Model predictions with likely future climate 

scenarios indicate continued diminishing snowfall, 

less snowpack, and less of a role for snowmelt 

in the hydrologic cycle of temperate regions of 

the Northern Hemisphere (IPCC, 2007). 

These patterns may affect the cycling of sulfur 

and nitrogen pollutants through ecosystems and 

watersheds by impacting episodic acidification, soil 

freezing, and the flows and temperatures in surface 

waters. When snow melts, surface-water chemistry 

is often at its most acidic and has its highest 

nitrate concentrations; therefore, a diminished 

snowmelt may lessen the episodic acidification 

of aquatic ecosystems (Moore et al., 1997). 

However, these predicted trends of diminishing 

snowmelt are uncertain, and an increase in midwinter rain has been noted in the eastern United 

States (Hodgkins et al., 2003). This increase in 

winter rain will also likely trigger strong episodic 

acidification, even when the rain falls on a 

diminished snowpack. Additionally, increases 

in large rainfall events have been observed and 

are predicted to increase with climate change in 

the twenty-first century (Murdoch et al., 2000). 

Increased episodic acidification from these large 

rainfall events may offset the predicted decreasing 

trend in snowmelt. In addition, diminished snow 

cover may change soil freezing patterns in winter, 

which are dependent on the magnitude of warming. 

Frequent freeze-thaw behavior in soils can increase 

nitrate leaching and raise nitrate concentrations 

in surface waters (Fitzhugh et al., 2003). 

Finally, earlier and diminished snowmelt may 

result in lower stream flows and warmer water 

temperatures in the later spring and summer. 

Warmer temperatures can result in increased 

growth rates. Increased aquatic growth can 

lower nitrate concentrations through biological 

uptake (Sommaruga-Wogarth et al., 1997); 

however, increased growth also can have negative 

impacts (e.g., hypoxia; see the section Effects 

of Atmospheric Nitrogen Deposition on Coastal 

Estuaries in this chapter for further discussion). 

Warmer temperatures and lower flows in streams 

may also have deleterious effects on coldwater fish species (Mohseni et al., 2003).

A nearly universal prediction from models of 

future climate is that droughts will increase in 

severity and duration, even in areas such as the 

eastern United States where overall increases in 

precipitation are predicted (IPCC, 2007). This 

forecast has strong implications for the impacts 

of acidification on aquatic ecosystems. When 

soils are dry, oxidized nitrogen and sulfur species 

accumulate and are released and flushed into local 

waterbodies upon re-wetting. Several studies have 

shown increased concentrations of nitrate and 

sulfate and decreased ANC values in surface waters 

when wet conditions return following drought 

periods. These impacts are especially apparent in 

wetland-influenced watersheds (Tipping et al., 

2003; Watmough et al., 2004; Schiff et al., 2005). 

This pattern of nitrogen and sulfur storage and 

release has further implications for the episodic 

acidification of waterbodies. Soils affected by 

acid deposition often contain levels of stored 

nitrogen and sulfur that are equivalent to decades 

of atmospheric deposition. Therefore, the 

aquatic ecosystem effects of severe, post-drought 

episodic acidification in lakes and streams and 

other surface waters may persist for decades after 

deposition levels have declined below critical 

loads (Tipping et al., 2003; Eimers et al., 2007). 

Rapid shifts from periods of dry to wet conditions 

are also expected to increase in the future. Increases 

in acidification driven by such climatic shifts have 

been shown to directly affect aquatic biological 

communities such as diatom algae (Faulkenham et 

al., 2003) and to alter physico-chemical variables 

that affect aquatic life, such as the penetration 

of ultraviolet radiation (Yan et al., 1996).

Nitrogen Availability has Important 

Implications for the Response of 

Ecosystems to Climate Change

One of the most uncertain and important 

ecosystem-related issues regarding the interaction 

of the atmospheric deposition of pollutants with 

changes in climate and CO2 levels concerns the role 

of nitrogen availability and its impact on growth 

and carbon sequestration. Growth in the majority 

of U.S. and global ecosystems is limited by nitrogen 

availability, suggesting that atmospheric nitrogen 
Chapter 3 | Effects of Acid Deposition on Ecosystems

66

deposition may be acting to increase growth and 

carbon sequestration in many ecosystems across the 

United States. However, in terrestrial ecosystems, 

excess atmospheric deposition of nitrogen also 

contributes to ecosystem acidification, calcium 

depletion in soils, loss of diversity, and excess 

runoff of nitrate, which can, in turn, lead to overenrichment in freshwater and estuarine ecosystems. 

Therefore, when considering atmospheric deposition 

of nitrogen, the potential “benefit” of increased 

growth and carbon sequestration must be weighed 

against the potential “costs” of resulting nutrient 

over-enrichment, decreased ecosystem services 

(e.g., biodiversity), and the promotion of invasive 

species (Fenn et al., 2003b). These issues are brought 

into sharp focus by studies that seek to determine the 

role of nitrogen in the present and future growth of 

terrestrial ecosystems (particularly forests) and the 

sequestration of carbon as climate and atmospheric 

CO2 concentrations continue to change.

One general conclusion is that net ecosystem 

growth in temperate and boreal forests has increased 

coincident with climate change in the United States, 

across Europe, and in parts of Asia over the past 

few decades (Goodale et al., 2002; Nemani et al., 

2003; Boisvenue and Running, 2006; Magnani 

et al., 2007). However, many local and regional 

exceptions to this generalization can be found 

(Korner, 2003). Another general conclusion is that 

atmospheric nitrogen deposition acts to increase 

net ecosystem growth and carbon sequestration in 

mid-latitude forests (Reich et al., 2006; Magnani et 

al., 2007), although the magnitude of this nitrogendriven carbon sink is likely much less than originally 

estimated by some studies (Sutton et al., 2008). 

Recent estimates of this sink are approximately 3% to 

21% of annual CO2 emitted globally from fossil fuel 

combustion (Churkina et al., 2009; Thomas et al., 

2009; Zaehle et al., 2010). The interactions of carbon 

and nitrogen are critical in controlling the magnitude 

of the terrestrial sequestration of a large amount of 

the anthropogenic (i.e., man-made) CO2 that would 

otherwise reside in the atmosphere and further 

alter the global climate. As CO2 in the atmosphere 

increases, terrestrial carbon sinks (especially forests) 

have been shown to increase; however, a point 

may be reached when nitrogen and other nutrients 

become limiting (Johnson, 2006; van Groenigen 

et al., 2006). The role of atmospheric nitrogen 

deposition in the global carbon cycle highlights 

the complex linkages among biogeochemical 

cycles and the important link between air 

pollutant deposition and global climate change.

Dissolved Organic Carbon—

Acid Deposition Interactions: 

A Case Study of Climate Change

Widespread increases in the concentrations and 

loads of dissolved organic carbon (DOC) in surface 

waters have been reported in the United States and 

Europe (Driscoll et al., 2003; Evans et al., 2006; 

Monteith et al., 2007). DOC is formed as organic 

matter decomposes and dissolves in water. DOC 

levels in surface waters are important because of the 

demonstrated link between atmospheric nitrogen 

and sulfur deposition and DOC loads in these 

waters (Clark et al., 2010). DOC plays a pivotal 

role in human health concerns related to water 

supply and distribution through the formation 

of disinfection by-products (Escobar et al., 2001; 

Chow et al., 2003). Additionally, DOC plays an 

important role in a wide array of aquatic ecosystem 

effects and interactions, including light penetration, 

water temperature, thermal stratification, plankton 

growth, pH and acidification, and the transport 

of trace metals (Snucins and Gunn, 2000). 

A variety of causes have been offered to explain 

why DOC concentrations are changing in remote 

fresh waterbodies that are not greatly influenced by 

human land use. These causes include the following: 

•	 Decreasing	atmospheric	sulfur	deposition	

(Evans et al., 2006; Monteith et al., 2007);

•	 Climate	warming	(Worall	and	Burt,	

2007; Clair et al., 2008);

•	 Changes	in	precipitation	amount	(Hudson	

et al., 2003; Worrall et al., 2003);

•	 Changes	in	incident	solar	radiation	

(Hudson et al., 2003);

•	 Decreases	in	sea	salt	deposition	

(Monteith et al., 2007); and

•	 Chronic	inputs	of	atmospheric	nitrogen	

deposition (Findlay, 2005).
NAPAP Report to Congress: An Integrated Assessment

67

d 

CASTNET monitoring site in Rocky Mountain National 

Park, Larimer County, CO (photo courtesy of U.S. EPA).

Reduced levels of sulfur deposition appear to 

be playing an important role in increased DOC 

concentrations in many surface waters in North 

America (Monteith et al., 2007), and the strength

of this association is likely to be greatest where 

atmospheric deposition loads are greatest (Clark 

et al., 2010). Decreases in atmospheric sulfur 

deposition over the past 10–20 years have resulted

in increasing pH and decreasing ionic strength in 

many waters, both of which should increase the 

solubility of organic matter in soils and sediment 

and, therefore, the DOC levels in surface waters. 

These observed increases in DOC concentrations 

may simply reflect ecosystem recovery and a 

return to the “natural” levels of DOC that were 

present before widespread acid deposition acidifie

aquatic ecosystems. Also, some observed DOC 

increases that appear to be driven by drought 

occurrence may actually result from the drought 

conditions favoring oxidation of previously 

reduced sulfur compounds that often originate 

from atmospheric deposition (Clark et al., 2006).

Climate variation also probably partly explains 

increasing surface-water DOC trends in some 

regions (Hudson et al., 2003; Worrall et al., 

2003). Fundamentally, warmer soil and sediment 

temperatures should increase the decomposition 

rate of organic matter and DOC levels in waters. 

However, variations in moisture, nutrient 

availability, the availability of dissolved oxygen, 

and other variables would be expected to mask 

a simple DOC to temperature relationship 

(Giardina and Ryan, 2000). For example, the 

Arctic regions of North America have experienced 

some of the greatest warming trends on Earth. 

Climate warming is expected to significantly 

increase DOC concentrations and fluxes in Arctic 

surface waters (Clair and Ehrmann, 1998; Frey 

and Smith, 2005). However, data from the Yukon 

basin supports decreased DOC export, which 

may result from the conversion of soil carbon to 

CO2 in the active soil permafrost layer (Striegl et 

al., 2005). Also, research conducted in Canadian 

lakes has highlighted the important role of other 

climate-related factors (e.g., drought-wetting 

cycles, variations in solar radiation) on surface 

water DOC concentrations (Dillon and Molot, 

1997; Schindler, 1998; Hudson et al., 2003).

Regardless of the relative roles of climate change 

and changes in atmospheric deposition of sulfur 

and nitrogen in driving these recent patterns of 

increased DOC concentrations and loads in surface 

waters, this issue highlights the interplay of acid 

deposition and climate change. Increases in DOC 

concentrations generally result in decreases in ANC;

therefore, these increased DOC concentrations 

are likely slowing the rate of recovery of aquatic 

ecosystems, as inferred from measurements of ANC 

(Monteith et al., 2007). Because of the widespread 

importance of DOC for aquatic ecosystems, future 

monitoring and modeling efforts should continue 

to scrutinize and consider the interplay of climate 

and atmospheric deposition. Historically, the 

recognition and study of DOC trends was not 

much discussed or considered until data from 

long-term monitoring networks, such as the LTM 

and TIME programs in the United States and other 

programs across the world, showed the magnitude 

and widespread geographic nature of these trends.

Terrestrial ecosystems store large amounts 

of carbon and the potential for these 

ecosystems to sequester or release additional 

carbon will likely contribute to future 

climate change. Research indicates that 

atmospheric nitrogen deposition affects the 

ability of terrestrial ecosystems to sequester 

atmospheric carbon dioxide.
Chapter 3 | Effects of Acid Deposition on Ecosystems

68

Summary of Climate Change— 

Atmospheric Deposition Linkages

The results summarized here provide a strong 

scientific basis for joint consideration of climate 

change and air quality policies affecting atmospheric 

nitrogen and sulfur deposition. Topics considered 

under the umbrella of global climate change include 

past, ongoing, and future expected changes in a 

wide range of factors that affect ecosystems, such as 

warmer air temperatures, water availability, and the 

frequency of large magnitude events (e.g., droughts, 

floods, fires). The combination of patterns in the 

atmospheric deposition of sulfur and nitrogen with 

these global climate change factors has implications 

for the growth of ecosystems, expected shifts of 

species, and the acid/base chemistry of surface 

waters. Untangling the interactions among multiple 

biogeochemical cycles and the resulting ecosystem 

responses to changes in these cycles is complex 

and uncertain. For example, global climate change 

generally provides an additional set of interacting 

stressors in ecosystems that are negatively impacted 

by acid deposition (see Figure 3-3). However, 

climate change also may increase rates of chemical 

weathering in some locations and, therefore, 

ecosystems may be able to sustain greater amounts 

of sulfur or nitrogen deposition before experiencing

negative effects. Conversely, atmospheric nitrogen 

deposition may currently be limiting global 

warming through enhanced ecosystem carbon 

sequestration in U.S. regions that are only lightly to 

moderately impacted by air pollutant deposition. 

Figure 3-3. Climate change, acid deposition, and other stressors (prepared by USGS).

Atmospheric Deposition

Sulfur and Nitrogen

Ecosystem Effects—Acidication 

and Nitrogen Over-enrichment

• Species changes

• Temperature-sensitive 

 biogeochemical processes

• Changes in water cycle

• Carbon dioxide increases in

 atmosphere

• Links with nitrogen cycle

• DOC increases in waters

Climate Change Changes in the Carbon Cycle

Note: Although the rate of atmospheric deposition of sulfur and nitrogen are the principal drivers of the 

ecosystem effects discussed in this chapter, ongoing and future changes in climate and the carbon cycle 

are expected to interact with these ecosystem effects in ways that are currently difficult to predict. Ongoing 

research efforts are helping to unravel the role of climate change on ecosystems so that more accurate 

model forecasts can better constrain the role of climate and carbon.

Currently, the degree of uncertainty regarding 

many of the predicted effects of global climate 

change on ecosystems is high and often not well 

quantified. Environmental monitoring programs 

are helpful in understanding climate-change effects 

on ecosystems, but often are not adequate in 

geographic scope or in the number of parameters 

measured to fully understand these effects. For 

example, the leading programs that measure 

atmospheric deposition (NADP and CASTNET) 

and surface-water chemistry (LTM and TIME) 

generally do not measure many climatically relevant 

variables, such as snow-water equivalents and air 

and water temperature. Additionally, several of 

the papers reviewed here highlighted the critical 

need for more multi-factorial and interdisciplinary 

experiments to test the effects of varying air/soil/

water temperature, CO2 concentrations, and 
NAPAP Report to Congress: An Integrated Assessment

69

pollutant deposition levels on ecosystems. There 

is also a need for consideration of climate-change 

effects in acidification and nitrogen effects models, 

such as MAGIC and PnET-BGC, as they are used 

to provide predictions of the expected impacts of 

future air pollutant deposition scenarios. Work of 

this type is ongoing, but not yet widely available in 

the published literature. Evaluations of the ability 

of these models to incorporate changing climatic 

effects would also be helpful in providing feedback 

to discussions about future policies. Finally, past and 

future disturbances that are not necessarily climatechange driven (e.g., land disturbance, land use 

changes, invasive species, fire) can have ecosystem 

effects that are as great or greater than those 

derived from climate change and are also worthy 

of consideration in future scenario modeling.

Multi-Pollutant Interactions 

in Ecosystems

Throughout this chapter, the interacting effects of 

nitrogen and sulfur deposition have been discussed 

in reference to the acidification and nutrient overenrichment of ecosystems. In the previous section, 

the interacting effects of acid deposition and 

changes in the carbon cycle and climate change were 

discussed. There are also important interactions 

between acid deposition and the ecosystem effects 

of ozone and mercury, and these interactions 

will be discussed in this section. In recent years, 

interest has been increasing in development of 

air quality policies that target multiple pollutants 

(NRC, 2004; Brook et al., 2009; NARSTO, 2011).

This section focuses on the ecosystem aspects 

of two pollutants (i.e., ozone and mercury) for 

which clear links have been demonstrated.

Ozone Interactions with Atmospheric 

Sulfur and Nitrogen Deposition

Ozone is taken up by plants and, in high 

concentrations, can directly damage plant cell 

membranes, reduce the rate of photosynthesis 

and growth, and cause species shifts in forests 

(Karnosky et al., 2007). In short, ozone is yet 

another pollutant stressor that can interact with 

sulfur and nitrogen deposition and changes 

in climate and CO2 concentrations to affect 

ecosystem health (McLaughlin and Percy, 1999). 

Ozone concentrations and trends are highly 

variable across the globe and the United States, 

but leveling or slight decreases in concentrations 

in North America have been noted in recent 

years (Vingarzan, 2004; Oltmans et al., 2006). 

Trends in ozone concentrations are affected by 

trends in (1) the sources of ozone, such as fossil 

fuel use, (2) the principle atmospheric precursors 

of ozone (i.e., NOx, volatile organic compounds, 

and carbon monoxide), and (3) climatic factors, 

particularly summer stagnation episodes that 

favor the highest concentrations. Results from 

models that simulate likely twenty-first century 

climate, combined with chemical transport model 

outputs, indicate a likelihood of increased future 

ozone concentrations in the United States (Jacob 

and Winner, 2009). Additionally, tropospheric 

ozone (ozone in the lowest portion of the Earth’s 

atmosphere) provides the third-strongest warming 

influence of all greenhouse gases (Mickley et 

al., 1999), and recent work suggests that ozone 

also contributes indirectly to warming by 

slowing growth, and subsequently CO2 uptake, 

in terrestrial ecosystems (Sitch et al., 2007).

Few studies have simultaneously considered the 

ecosystem effects of ozone combined with sulfur 

and nitrogen deposition. The added complication of 

ongoing changes in the climate and the carbon cycle 

provide additional challenges for understanding 

patterns and predicting likely ecosystem impacts of 

ozone, sulfur, and nitrogen loading. The inability 

of current models to adequately consider the 

simultaneous effects of these pollutants has been 

noted (Aber et al., 2001; Dewar et al., 2009). 

Despite the lack of adequate models and sufficient 

experiments to understand multi-factorial variations 

of these pollutants, evidence supports a joint, 

interacting role of ozone and nitrogen deposition in 

a variety of ecosystem effects, including sensitivity 

to insects and pathogens; frost sensitivity; drought; 

and fire (Bytnerowicz et al., 2007). Thresholds 

of nitrogen deposition and ozone concentration 

likely exist, below which ecosystem effects cannot 

be detected and at which benefits such as increased 

ecosystem growth and carbon sequestration are 

likely. However, the exact levels of these thresholds 

are not well known, nor are the tipping points above 
Chapter 3 | Effects of Acid Deposition on Ecosystems

70

Great Smoky Mountains National Park protects one of 

the last wild trout habitats in the eastern United States 

(photo courtesy of the NPS).

which negative ecosystem effects sharply increase 

(Aber et al., 2001). Better identification of such 

patterns in ecosystem effects could provide vital 

information for use in future air quality policies. 

Mercury Interactions with 

Atmospheric Sulfur and 

Nitrogen Deposition

Although most surface waters have very low 

concentrations, mercury is of environmental 

concern because it accumulates in living cells 

and is biomagnified in aquatic and terrestrial 

food webs to the extent that elevated mercury 

concentrations are present in fish throughout the 

United States (Scudder et al., 2009). Mercury is 

a potent neurotoxin, and human health warnings 

that suggest limiting the consumption of certain 

fish due to high mercury levels are widespread in 

the United States (U.S. EPA, 2007). Atmospheric 

sulfur and nitrogen are closely linked to mercury 

because emissions from coal burning are a major 

source of all three of these pollutants. Moreover, 

atmospheric sulfur and nitrogen deposition 

can affect the cycling and bioaccumulation 

of mercury in ecosystems, primarily because 

of the interactions of mercury and sulfur.

The mercury found in the tissue of biota at high 

ecosystem trophic levels (e.g., predatory fish 

and birds) is dominantly in the methyl form 

(methylmercury), and this is the dominant form 

found in humans as well. Methylation is the 

biogeochemical process by which mercury is 

converted to methylmercury, and most methylation 

in ecosystems is believed to be carried out by the 

same bacteria that convert sulfate to sulfide forms. 

This process tends to proceed in environments 

(e.g., wetlands, lake-bottom sediments) where 

sulfate is introduced in runoff and oxygen is 

absent. Several studies have shown that additions 

of sulfate increase methylmercury concentrations 

in waters, and by inference, would be expected to 

also increase mercury levels in biota (Branfireun 

et al., 1999; Gilmour et al., 1992; Jeremiason 

et al., 2006). Similar studies have shown that 

decreases in sulfate concentrations can likewise 

decrease methylmercury concentrations in waters 

and fish tissue, even in the absence of any changes 

in atmospheric mercury deposition rates (Hrabik 

and Watras, 2002; Drevnick et al., 2007). 

The findings from these studies suggest that 

improvements in ecosystem health, with respect 

to mercury bioaccumulation levels, might be 

achieved by simultaneously decreasing sulfur and 

mercury deposition levels. These findings also 

suggest that a multi-pollutant policy, considering 

both mercury and sulfur, could be used strategically 

to reduce mercury bioaccumulation in many 

environmental settings. In addition to direct 

links between methylmercury and sulfate, studies 

have found that pH and mercury levels in fish 

tissue are inversely related in many regions. This 

suggests that ecosystem acidification and mercury 

bioaccumulation are also linked, and a variety of 

mechanisms have been suggested to explain this 

relationship (Kamman et al., 2004; Wiener et al., 

2006; Munthe et al., 2007; Scudder et al., 2009). 

Another recent study shows that high nitrate 

concentrations in waters may act to suppress the 

formation of methylmercury when environmental 

conditions might otherwise favor high rates of 

methylation (Todorova et al., 2009), but this type 

of interaction has not yet been demonstrated in a 

natural ecosystem where atmospheric deposition 

is the principal source of nitrogen and sulfur.
CHAPTER 4

Beyond Title IV: Ecological Impacts 

of Further Emission Reductions

Title IX of the 1990 CAAA requires NAPA

to report quadrennially on “the reduction 

in deposition rates that must be achieved in

order to prevent adverse ecological effects”

(Public Law 101-549-Nov. 15, 1990). NAPAP

presented a working definition of “adverse 

ecological effects” in the 1996 NAPAP 

RTC (NSTC, 1998) based on the intent of 

Congress, as expressed in the 1990 CAAA 

and shaped by other relevant environmental

statutes (i.e., Comprehensive Environmental

Response, Compensation and Liability Act, 

and the Clean Water Act) and associated 

regulations. The definition is as follows: 

P 

Adverse ecological effects: any injury 

(i.e., loss of chemical or physical 

quality or viability), to any ecological 

or ecosystem component, up to 

and including at the regional level, 

over both long- and short-terms.

At a Glance: Modeling Results 

•	 Additional	emission	reductions	are	still	

needed for ecological recovery, especially in 

acid-sensitive areas.

•	 Reduction	of	SO2	and	NOx to below 1.10 

and 0.44 million tons/year from the power 

sector, respectively, would dramatically 

reduce the number of lakes classified as 

waters of acute concern.

•	 Reduction	of	SO2 to 0.44 million tons/

year from the power sector would result in 

nearly full protection of lakes in the Northeast and the Adirondack Mountains.

•	 Some	lag	between	emission	reductions	and	

ecological recovery is expected in certain 

areas, especially streams in the southeastern United States.

Adverse impacts to ecological processes or 

ecosystem components include the results of 

reductions in ANC and pH and increases in 

aluminum concentrations in a lake or stream; 

loss of fish and other biota; loss of important 

nutrients, such as calcium, from forest soils; and 

increased susceptibility of trees to pests, disease, 

and winter temperatures. These, in turn, lead 

to decreasing forest growth and forest dieback. 

Adverse ecological impacts also include the 

effects of nitrogen saturation in forests; effects in 

alpine lakes; coastal eutrophication as a result of 

atmospheric deposition; reductions in biodiversity; 
Chapter 4 | Beyond Title IV: Ecological Impacts of Further Emission Reductions

72

fire regime shifts; and injuries to plants as a result 

of ozone exposure. Other effects of impaired air 

quality addressed by NAPAP (e.g., materials, 

visibility, human health) are not considered here 

because this chapter focuses on ecological effects.

Deposition levels that correlate with a “threshold” 

of adversity are scientifically complex and can 

be difficult to establish because most biological 

responses to changes in acid–base chemistry occur 

along a continuum, with no single value or set of 

chemical concentrations that represents a threshold 

for “significant adverse biological effects” (Bulger 

et al., 1999; Sullivan et al., 2008; U.S. EPA, 

2008b, 2009f). Therefore, this report describes 

ecosystem responses along a continuum, enabling 

decision makers to determine levels of acceptable 

risk. The analysis in this chapter summarizes the 

results into categories of acidification concerns, and 

these categories describe the path that lakes and 

streams follow in recovery from acidic precipitation 

(Figure 4-1), as indicated by changes in water 

chemistry and inferred biological response. 

Figure 4-1. Stages of recovery for acidic lakes and streams (prepared by U.S. EPA).

Waters of Acute Concern

Water acidic all the time;

sensitive organisms

cannot survive

Waters of Low Concern

Recovery of water

chemistry; even sensitive

organisms can survive

Waters of Moderate Concern

Signicant recovery, but water

still acidied at times; sensitive

organisms may survive

Waters of Elevated Concern

Water still acidied seasonally

or after storms; sensitive

organism richness still low

Note: See text box titled “Acid Neutralizing Capacity” in Chapter 2 for a full discussion of the ecological effects of acidification. 

Waters of acute concern are defined as having ANC less than 0 μeq/L. 

Waters of elevated concern are defined as having ANC from 0 to less than 50 μeq/L. 

Waters of moderate concern are defined as having ANC from 50 to less than 100 μeq/L. 

Waters of low concern are defined as having ANC greater than or equal to 100 μeq/L.

Recent Assessments

Although the definition of a “threshold” is complex

a significant amount of research in the past decade 

indicates that ecosystems continue to be affected by

, 

acid deposition. Two recent syntheses of the science 

found that, for the near term, sulfur deposition 

still is the primary source of acidification in most 

sensitive areas of eastern North America (U.S. 

EPA, 2008b, 2009f). However, these syntheses 

also found that nitrogen deposition, particularly 

ammonia, is playing a greater role in both shortand long-term acidification of lakes and streams as 

ammonia is nitrified to nitrate, which can acidify 

soils and surface water when the amount exceeds 

biological uptake. Under certain deposition levels, 

sulfur and nitrogen could have approximately 

equal roles in surface water acidification. Decreases 

in emissions of SO2 and increases in the level of 

nitrogen saturation of forest soils have contributed 

to the increasing role of nitrogen in surfacewater acidification in the eastern United States.

In the period since the 2005 NAPAP RTC, various 

studies (Warby et al., 2005; Burns et al., 2006; 

Driscoll et al., 2007; Sullivan et al., 2008; Lawrence 

et al., 2008; Robinson et al., 2008) concluded that 

current acid deposition loads are causing ecological 

damage in sensitive lakes and streams in the eastern 

United States. The results presented in Chapter 2 

of this report, as well as studies focused on specific 

regions of the country, demonstrate improving 

water quality in most of the Adirondack Mountains, 

New England, and the Northern Appalachian 

Plateau. However, many lakes and streams still 

remain impacted by acid deposition under current 

deposition levels (U.S. EPA, 2009a, 2009d). For 

example, the extent of stream acidification in the 

western Adirondack Mountains remains high. The 

streams sampled in a recent study showed that 66%, 

or 718 km, of streams are prone to acidification 

and likely have levels of acidity harmful to their 

biota. Of the 66% of streams found to be prone to 

acidification, about half were likely to be chronically 

acidified (i.e., base cation surplus < 0 μg/L), with 
NAPAP Report to Congress: An Integrated Assessment

73

the other 50% episodically acidified (i.e., base 

cation surplus > 25 μg/L) (Lawrence et al., 2008).

A major assessment of acid deposition and its 

effects in the Adirondack Mountains found that 

full implementation of the 1990 CAAA will 

result in only modest recovery in northeastern 

lakes and streams impacted by acid deposition 

(Sullivan et al., 2006a). However, based on 

model projections, this study also concluded that 

aggressive reductions of SO2 and NOx emissions 

from power generation of up to 70% and 50%, 

respectively, beyond 2001 levels, would allow 

for chemical recovery (i.e., become nonacidic) 

of most lakes in the Adirondack Mountains. 

Researchers in the southern Appalachian Mountains 

concluded that streams in this region are still 

threatened by acid deposition (Sullivan et al., 

2008). In particular, many brook trout streams 

in Virginia are still vulnerable to acidification at 

current deposition levels (Cosby et al., 2006). 

Based on model projections from two recent 

studies (Cosby et al., 2006; Sullivan et al., 2008), 

researchers concluded that further reductions of 

sulfate deposition beyond levels achieved by the 

Title IV SO2 emission reductions are necessary 

to prevent further acidification of southern 

Appalachian Mountain streams and to allow 

currently impacted streams to recover in the region. 

Cosby et al. (2006) concluded that a moderate 

reduction of sulfur and nitrogen deposition of 50% 

and 22%, respectively, beyond Title IV produced 

a small improvement in stream water quality 

over the long term (by 2100) relative to current 

conditions for brook trout streams in Virginia. 

A more stringent scenario of 62% reduction for 

sulfur and 30% reduction for nitrogen deposition 

produced additional improvements in stream water 

quality over the long term (by 2100), although these 

reductions still did not return stream water quality 

to preacidification conditions (Cosby et al., 2006). 

In sensitive streams in North Carolina, Tennessee, 

and South Carolina, Sullivan et al. (2008) modeled 

similar emission reductions, as did Cosby et al. 

(2006), and found little future improvement in 

stream conditions resulting from moderate and 

aggressive emission controls (Sullivan et al., 2008).

Analysis of the Environmental 

Impact of Further Air 

Emission Reductions

In response to the CAAA Title IX statutory 

requirement to identify deposition rates that would 

prevent adverse ecological effects, this chapter 

analyzes several emission-reduction scenarios that 

broadly bound the range of reductions presented 

in the literature discussed above and are similar to 

scenarios modeled in past NAPAP reports. These 

results provide an indication of the environmental 

improvements that would be expected from 

additional emission reductions from sources 

affected by Title IV and other emission-reduction 

programs. These environmental improvements 

do not constitute full recovery in all areas of acidsensitive forests, lakes, or streams that have been 

impacted by acid deposition. They do provide, 

however, an indication of the scope and magnitude 

of the impact of emission reductions on deposition 

levels and on acid-sensitive ecosystems (Figure 4-2). 

Acid-sensitive ecosystems are generally characterized 

by a low ability to buffer against acidification due 

to hydrologic, geologic, and soil characteristics. 

Other sensitive ecosystems, such as estuaries and 

western U.S. alpine lakes, would also benefit from 

reductions in nitrogen deposition, but effects on 

those ecosystems are not analyzed here. In addition, 

climate change may alter the response of lake and 

stream water quality to declining acidic deposition 

(see Chapter 3 for a discussion of climatic change 

and surface-water recovery). However, this modeling 

analysis, like most previous scientific studies of 

surface-water recovery from acidification, assumes 

a constant influence of climate over the simulation 

period. While the information presented in this 

chapter is useful to inform future actions, the 

analysis presented here focuses solely on ecological 

recovery. Other aspects of implementing these 

emission-reduction scenarios, such as the costs 

and additional benefits (e.g., human health), 

were not analyzed and are beyond the scope of 

this assessment. Additional information would be 

important to inform future actions, including the 

costs and other impacts of emission reductions from 

the power sector and other sectors and the value 

that the public places on further improvements 

to the environment and human health.
Chapter 4 | Beyond Title IV: Ecological Impacts of Further Emission Reductions

74

Scenarios Analyzed

This modeling effort analyzed several scenarios 

representing emission changes from stationary 

sources within the power sector for the purpose of 

evaluating ecological recovery. Results are presented 

for total sulfur and total nitrogen deposition because 

these pollutants are the primary components of 

acid deposition and the cause of anthropogenic 

surface-water acidification. This analysis focuses 

on reductions from the power generation sector 

(i.e., EGUs); however, many other sources also 

emit sulfur and nitrogen pollutants that contribute 

to acid deposition. For example, sources outside 

of the power generation sector are projected to 

emit approximately 52% of the SO2 and 84% of 

the NOx emitted in 2020 under the Base Case 

scenario described below (U.S. EPA, 2006c). 

Acid-Sensitive Areas

Western United States

New England

Northeastern Coastal Plain

Ouachita Mountains 

Upper Midwest

Southern Appalachians

Florida Highlands

Central Appalachians and Northern Appalachian Plateau

Adirondack Mountains

Figure 4-2. Map of acid-sensitive ecosystems in the United States (prepared by U.S. EPA).

The following analysis compares emission and 

deposition levels under a Base Case scenario with 

three other sensitivity scenarios of various levels 

of additional reductions. The Base Case scenario 

consists of currently implemented programs and 

programs that were finalized as of spring 2005. 

These programs include Title IV, CAIR, and Tier II 

and Heavy Duty Diesel SO2 and NOx reductions. 

The analysis employs models and analytical 

tools that are widely used and peer reviewed. 

EPA’s Response Surface Model (RSM) was used 

to estimate future deposition loads (U.S. EPA, 

2006c). RSM is based on an air quality modeling 

approach known as meta-modeling that aggregates 

numerous individual Community Multiscale 

Air Quality model (CMAQ) simulations into a 

multidimensional air quality response surface. The 

Integrated Planning Model (IPM) was used to 
NAPAP Report to Congress: An Integrated Assessment

75

estimate the power sector emissions from EGUs. 

As with any detailed analysis of complex scenarios, 

the results presented in this report are subject to 

uncertainties concerning emission estimates, air 

quality modeling, deposition projections, and the 

impact of emission reductions on ecological systems. 

See the text box titled Modeling Tools Used in this 

Analysis for a more detailed description of the air 

quality and emission models used in this analysis.

The scenarios described below represent emission 

levels for each pollutant in 2020. Because the Base 

Case scenario includes implementation of CAIR, 

resulting in significant SO2 and NOx emission 

reductions beyond levels achieved by Title IV 

alone, the levels of SO2 and NOx modeled for 

the other future scenarios include additional 

reductions from the CAIR SO2 and NOx cap levels. 

The projected emissions used for the air quality 

modeling in 2020 are somewhat higher than the 

cap levels for all scenarios as a result of the early 

reductions and allowance banking predicted by the 

emission model (i.e., IPM). For the mobile sectors, 

the growth and controls are calculated together 

by the MOBILE6 and NONROAD models 

(U.S. EPA, 2006c). No controls were applied 

to the Canadian, Mexican, or offshore emission 

inventories. Additional details on the scenarios are 

summarized in Figure 4-3 and presented below. 

•	 Base Case Scenario (2020). This scenario 

includes rules that were finalized as of spring 

2005, including CAIR and the Non-Road Diesel 

Rule. It does not include new or anticipated 

actions under the Clean Air Act or other 

emission reductions that would be necessary to 

attain and maintain the fine particle and ozone 

NAAQS for which states are required to submit 

SIPs or to achieve regional haze-reduction goals. 

•	 Scenario A. This scenario includes an additional 

60% reduction in SO2 emissions from the power 

generation sector beyond the Base Case scenario, 

resulting in an annual SO2 emission level of 

1.75 million tons/year in 2020. It also includes 

national annual NOx emission reductions from 

the power generation sector of 68% beyond 

the Base Case scenario, or an annual emission 

level of 0.7 million tons/year of NOx in 2020. 

•	 Scenario B. This scenario includes an additional 

75% reduction in SO2 emissions from the 

power generation sector beyond the Base Case 

scenario, equaling annual SO2 emissions of 

1.10 million tons/year in 2020. It also includes 

national annual NOx emission reductions 

from the power generation sector of 80% 

beyond the Base Case scenario, or annual NOx

emissions of 0.44 million tons/year in 2020. 

•	 Scenario C. This scenario is roughly equivalent 

to elimination of SO2 emissions from the power 

generation sector. It includes an additional 90% 

reduction in SO2 emissions from the power 

generation sector beyond the Base Case scenario, 

resulting in 0.44 million tons SO2/year in 

2020, and a 50% reduction in SO2 emissions 

from nonpower generation sources (e.g., 

industrial boilers). The scenario also includes 

national annual NOx emission reductions of 

approximately 80% beyond the Base Case 

scenario for the power sector, resulting in annual 

NOx emissions of 0.44 million tons/year in 2020.

0

0.5

1

1.5

2

2.5

3

3.5

4

4.5

5

SO2 & NOx Emissions (million tons/yr)

SO2

NOx

Scenario in 2020

Base Case A B C

Figure 4-3. Projected annual SO2 and NOx emission 

levels in 2020 for the Base Case scenario and Scenarios 

A through C. Emission levels are for the continental 

United States (prepared by U.S. EPA).
Chapter 4 | Beyond Title IV: Ecological Impacts of Further Emission Reductions

76

Modeling Tools Used in this Analysis

The Integrated Planning Model (IPM) is used to analyze the projected impact of environmental policies 

on the electric power sector in the 48 contiguous states and the District of Columbia. IPM is a multiregional, 

dynamic, deterministic linear programming model of the U.S. electric power sector. It provides forecasts of 

least-cost capacity expansion, electricity dispatch, and emission-control strategies for meeting energy demands and environmental, transmission, dispatch, and reliability constraints. IPM can be used to evaluate the 

cost	and	emission	impacts	of	proposed	policies	to	limit	emissions	of	SO2,	NOx,	CO2, and mercury from the 

electric power sector. IPM provides the EGU emission inventory used as an input to CMAQ.

Air Quality and Deposition Models—Air quality models use mathematical and numerical techniques 

to simulate the physical and chemical processes that affect air pollutants as they disperse and react in the 

atmosphere. Using inputs of meteorological data and source information such as emission rates, these models 

are designed to characterize primary pollutants that are emitted directly into the atmosphere and, in some 

cases, secondary pollutants that are formed as a result of complex chemical reactions within the atmosphere. 

•	 The Community Multi-scale Air Quality 

(CMAQ) model is a three-dimensional, regional 

grid-based air quality model designed to simulate 

air quality and deposition over the contiguous 

United States for a period of 1 year. The CMAQ 

model includes state-of-the-science capabilities 

for conducting urban- to regional-scale simulations of multiple air quality issues, including 

tropospheric ozone, fine particles, toxics, acid 

deposition, and visibility degradation. The CMAQ 

model is a publicly available peer-reviewed, stateof-the-science model supported by the Community Modeling and Analysis System (CMAS) 

Center (http://www.cmascenter.org).

•	 The Response Surface Model (RSM) is based 

on an approach known as air quality meta-modeling that statistically links pollution emissions and 

air quality/deposition derived from other models, 

such as the CMAQ model. The RSM aggregates 

numerous individual air quality modeling simulations from the CMAQ model to produce a multidimensional air quality response surface, which 

can be used to predict how emission changes 

affect air quality. The RSM approach allows for 

the evaluation of how emission-control scenarios 

improve air quality across the United States. The 

RSM model was used in support of the Regulatory Impact Analysis (RIA) for the proposed 

NAAQS for PM2.5 (U.S. EPA, 2006b).

Water/Watershed Modeling—The Model of Acidification of Groundwater in Catchments (MAGIC) 

was developed to estimate acidification of lakes and streams in response to sulfur and nitrogen deposition 

(Cosby et al., 1985a,b,c; U.S. EPA, 2006c). MAGIC was the principal model used by NAPAP to estimate future damage and recovery to lakes and streams in the eastern United States in the 1998 and 2005 integrated assessments (NSTC, 1998 and 2005). The model simulates soil solution chemistry and lake and stream 

chemistry to predict the monthly and annual average concentrations of the major ions in these waters. 

MAGIC consists of (1) a submodel in which the concentrations of major ions are assumed to be governed 

by simultaneous reactions involving sulfate adsorption, cation exchange, dissolution-precipitation-speciation 

of aluminum, and dissolution-speciation of inorganic carbon; and (2) a mass balance submodel in which 

the flux of major ions to and from the soil is assumed to be controlled by atmospheric inputs, chemical 

weathering, net uptake and loss in biomass, and losses to runoff. Nitrogen retention and loss is modeled in 

two ways. The simpler method (used for the majority of the analysis presented in this report) assumes that 

the percentage of nitrogen deposition retained by the soil remains constant over time. The more complex 

approach simulates nitrogen saturation and links net immobilization of nitrogen to the carbon/nitrogen 

ratio of the soil organic matter pool. Both require specification of net nitrogen uptake in vegetation, rate 

of denitrification, and nitrogen fixation. At the heart of MAGIC is the size of the pool of exchangeable base 

cations in the soil. As the fluxes to and from the pool change over time because of changes in atmospheric 

deposition, the chemical equilibria between soil and soil solution shift to give changes in surface water 

chemistry. MAGIC provides a widely accepted tool for modeling the response of lake and stream chemistry to sulfur and nitrogen deposition.
NAPAP Report to Congress: An Integrated Assessment

77

Total Sulfur Deposition

(kg/ha)

< 0.5

0.5–2.5

2.5–5.0

5.0–7.5

7.5–10.0

10–12.5

12.5–15

> 15

Scenario A

Base Case

Scenario B

Scenario C

Scenario B

Changes in Sulfur Deposition

All future control scenarios modeled in this 

analysis are projected to lead to significant regional 

reductions in sulfur deposition as compared to 

projected conditions under the Base Case scenario 

in 2020 (Figure 4-4). Modeling of the Base Case 

scenario indicates that implementation of current 

rules in 2020 is expected to reduce sulfur deposition 

from 2010 levels by greater than 5 to 10 kilograms/

hectare/year (kg/ha/yr) for much of the eastern 

United States. Under Scenario A in 2020, reductions 

in sulfur deposition of 10% to 20% beyond 2020 

Base Case levels would be found in much of the 

Plains states and Northeast, with reductions of 30% 

to 45% in the Midwest, Northern Appalachian 

Plateau, central Appalachians, and southern 

Appalachians stretching from Indiana, Ohio, and 

Pennsylvania to Georgia and Alabama. The western 

states, from the Rocky Mountains west, would see 

a mixture of reductions and increases in deposition 

of about 10% under Scenario A (Figure 4-5).

Figure 4-4. Projected change in annual total sulfur deposition from Base Case in 2020 with additional SO2 and NOx 

emission reductions from the power generation sector and industrial boilers (prepared by U.S. EPA). 
Chapter 4 | Beyond Title IV: Ecological Impacts of Further Emission Reductions

78

Total Sulfur Deposition

(Percent change)

10 to 20

0 to 10

-10 to 0

-20 to -10

-30 to -20 

< -30

Scenario C

Scenario B

Scenario A

Figure 4-5. Projected percentage change in annual total sulfur deposition from Base Case in 2020 with additional 

SO2 and NOx emission reductions from the power generation sector and industrial boilers (prepared by U.S. EPA).

Scenario B extends the area of largest reduction 

in sulfur deposition southward and westward. 

For example, under Scenario B, the central 

Appalachians, Northern Appalachian Plateau, 

southern Appalachians, and Piedmont regions 

would all experience reductions in sulfur 

deposition beyond levels achieved by the 2020 

Base Case scenario. Scenario C would extend 

the area with the greatest reductions in sulfur 

deposition beyond the 2020 Base Case scenario 

northward to Maine and westward through the 

Midwest across the Plains states to the Rocky 

Mountains (Figure 4-5), with some areas in the 

eastern United States experiencing reductions up 

to 60%. These reductions are expected to provide 

ecological benefits to these acid-sensitive regions, 

but will not be sufficient to allow full recovery. 

Many of the most sensitive sites within these acidsensitive regions will not likely gain full recovery 

within the timeframe modeled in this analysis.
NAPAP Report to Congress: An Integrated Assessment

79

Changes in Nitrogen Deposition

Modeling of the 2020 Base Case scenario indicates 

that, in 2020, rules currently in effect are expected 

to achieve reductions in nitrogen deposition of 

2.5 to 5 kg/ha/yr beyond 2010 levels across the 

United States (Figure 4-6). Under Scenario A, 

additional reductions in deposition of 10% to 

20% beyond the levels seen under the 2020 Base 

Case scenario would occur in much of the central 

Appalachians and the eastern Ohio River Valley 

stretching from northern Pennsylvania to southern 

Virginia (Figure 4-7). Similar levels of deposition 

reductions would be found in the Four Corners 

region (i.e., Arizona, Colorado, New Mexico, Utah) 

of the western United States and along the Front 

Range of the Rocky Mountains. The model results 

suggest a slight increase in deposition in east Texas, 

western Arkansas, and eastern North Carolina. In 

addition, west of the Rocky Mountains, some areas 

would experience nitrogen deposition increases 

of up to 10%, including areas in Montana and 

Idaho and the Sierra Nevada and San Bernardino 

mountains of California. Increases in these regions 

are a result of higher NOx emissions from other 

sources, such as non-EGU industrial sources. 

The rest of the country would see additional 

reductions in deposition of up to 10% beyond the 

levels seen under the 2020 Base Case scenario. 

Total

Nitrogen Deposition

(kg/ha)

< 0.5 

0.5–2.5

2.5–5

5.0–7.5

7.5–10

10.0–12.5

> 12.5

Scenario A

Base Case

Scenario B

Note: Because Scenarios B and C have 

the same level of NOx emission 

reductions, Scenario C is not shown.

Figure 4-6. Projected change in annual total nitrogen deposition from Base Case in 2020 with additional SO2 and 

NOx emission reductions from the power generation sector (prepared by U.S. EPA). 
Chapter 4 | Beyond Title IV: Ecological Impacts of Further Emission Reductions

80

Scenario B extends the area of 10% to 20% 

additional reductions in nitrogen deposition to the 

upper Midwest and parts of the Northeast, along 

the Colorado Front Range, and in the Four Corners 

region, with regions in West Virginia experiencing 

up to 30% reductions. Large portions of the United 

States, including the Plains states, Midwest, and 

Western states, will experience minimal reductions 

in nitrogen deposition of 0% to 10%. However, 

areas in Montana, Idaho, and California will 

continue to have increases in deposition of up 

to 10% (Figure 4-7), as a result of higher NOx

emissions form other sources. Emission levels in 

Scenario B would also lead to significant additional 

reductions in nitrogen deposition compared to the 

2020 Base Case scenario in sensitive ecosystems 

still experiencing water quality or forest health 

problems as a result of acidification. For example, 

the Adirondack Mountains would receive a 

20% reduction in deposition under Scenario B 

as compared to the 2020 Base Case scenario.

NOx emissions are the same in Scenarios B 

and C. Although particulate chemistry would 

indicate some change in nitrogen deposition 

based on changes in sulfur emissions, Scenario 

C showed no significant differences in nitrogen 

deposition compared to Scenario B.

Scenario B

Scenario A

Total Nitrogen Deposition 

(Percent change)

10 to 20

0 to 10

-10 to 0

-20 to -10

< -20 

Note: Because Scenarios B and C have 

the same level of NOx emission 

reductions, Scenario C is not shown.

Figure 4-7. Projected percentage change in annual total nitrogen deposition from Base Case in 2020 with additional 

SO2 and NOx emission reductions from the power generation sector (prepared by U.S. EPA).
NAPAP Report to Congress: An Integrated Assessment

81

Additional Ecological Impacts of 

Decreases in Nitrogen Deposition

In addition to acidification, nitrogen deposition 

can contribute to the eutrophication of estuaries 

and nutrient enrichment of forests. For example, 

based on the results of the Response Surface 

Model (RSM), the Chesapeake Bay Estuary would 

receive an approximate 10% to 20% reduction in 

nitrogen deposition from the power generation 

sector under Scenarios B and C as compared 

to the 2020 Base Case scenario. Additionally, 

forests in the Front Range region of Colorado 

would receive approximately a 20% reduction in 

nitrogen deposition from the power generation 

sector in 2020, with reductions over 30% in some 

areas, under Scenarios B and C as compared to 

the Base Case scenario. Additionally, forests in the 

Front Range region of Colorado would receive 

approximately a 20% reduction in nitrogen deposition in 2020, with reductions over 30% in some 

areas, under Scenarios B and C as compared to 

the Base Case scenario. 

Regional Ecosystem Implications 

of Changes in Sulfur and 

Nitrogen Deposition 

The projected sulfur and nitrogen deposition 

displayed in Figure 4-4 and 4-6 represents the 

amount of total sulfur and nitrogen deposition 

a given region will receive in 2020. As described 

in Chapter 3 of this report, sulfur and nitrogen 

deposition has the potential to cause ecological 

harm, including acidification, to sensitive 

ecosystems. Not all deposition loads will cause 

ecological harm, as some ecosystems have the 

ability to neutralize acids that are deposited from 

the atmosphere. However, some ecosystems have a 

limited ability to neutralize even low levels of acidic 

deposition to surface waters and soils. When these 

ecosystems exceed their ability to neutralize acidic 

deposition, surface waters and soils acidify to the 

point where aquatic and terrestrial life is harmed. 

The degree to which an ecosystem will be harmed is 

controlled by the level of ecosystem sensitivity and 

deposition load. Moreover, patterns of ecosystem 

sensitivity vary by region of the country. Typically, 

in the eastern United States when deposition loads 

of sulfur and nitrogen are greater than 12 kg/ha/

yr and 8 kg/ha/yr, respectively, widespread effects 

of acidification occur in surface waters. When 

deposition loads are below 5 kg/ha/yr for both 

sulfur and nitrogen, the most sensitive ecosystem 

elements are generally protected. While any level of 

deposition below 12 kg/ha/yr of sulfur and 8 kg/

ha/yr of nitrogen provides additional protection, 

the most sensitive ecosystems are likely still strongly 

impacted (Aber et al., 2003; Brookshire et al., 

2007). In special cases, particularly in western 

states, considerably lower levels of combined sulfur 

and nitrogen deposition may acidify headwater 

streams. In the Rocky Mountains, headwater 

streams can become episodically acidified during 

snowmelt when deposition loads are as low as 

2 kg/ha/yr for combined sulfur and nitrogen. 

Nitrogen deposition also contributes to nutrient 

over-enrichment of ecosystems. Nitrogen is a 

nutrient required by living systems, and so the 

level of nitrogen in an ecosystem often limits 

the system’s productivity (i.e. the growth rate 

of plants and other organisms in the system). 

Because nitrogen is an important nutrient, 

the majority of nitrogen deposition is used or 

stored by ecosystems to allow for increased 

productivity. Nitrogen over-enrichment may cause 

eutrophication and lead to changes in species 

composition. The degree to which an ecosystem 

will be over enriched, and potentially harmed, is 

controlled by the ability of the ecosystem to use the 

additional nitrogen deposition, but still maintain 

a healthy and diverse ecological community. 

•	 For	many	Rocky	Mountain	ecosystems,	

nitrogen deposition levels as low as 1.5 

kg/ha/yr correspond to shifts in aquatic 

species composition, such as lake plankton, 

and levels as low as 4 kg/ha/yr can cause 

similar shifts in alpine plant communities 

(Baron, 2006; Bowman et al., 2006). 

•	 In	many	areas	of	the	far	western	United	States,	

native plant communities such as coastal sage 

scrub, desert scrub, and annual grasslands 

are adapted to low-nitrogen environments. 

When nitrogen loads to these habitats 
Chapter 4 | Beyond Title IV: Ecological Impacts of Further Emission Reductions

82

exceed 5–8 kg/ha/yr, the native plant and 

soil communities can be replaced by invasive 

species that are better adapted to more nitrogenrich environments (Fenn et al., 2003a). 

•	 For eastern and northern forests, the amount of

nitrogen deposition generally harmful to forest 

ecosystems is less clear. Studies have shown that 

nitrogen deposition exceeding 8–26 kg/ha/yr led 

to increased surface water nitrate leaching and 

declines in growth and survivorship of sensitive 

tree species (Aber et al., 2003; McNulty et al., 

2005). However, other studies based on modeling 

find that negative effects could occur in eastern 

and northern forests when nitrogen deposition is 

as low as 3–7 kg/ha/yr (Pardo and Duarte, 2007).

Changes in the Ecological 

Condition of Lakes and Streams

The magnitude of emission reductions influences 

both the amount of recovery from acidification 

and the rate at which recovery occurs. The rate of 

recovery is also influenced by the geological and 

ecological characteristics of the lakes or streams 

in the area. The path of ecological recovery from 

acidification in lakes and streams is displayed in 

Figure 4-1. In this analysis, the ecological response 

to emission reductions of lakes in the Adirondack 

Mountains and the Northeast and streams in 

the Southeast (central and southern Appalachian 

Mountains) were modeled using MAGIC, which 

estimates acidification of lakes and streams in 

response to sulfur and nitrogen deposition (Wright 

et al., 2006) (see the text box, Modeling Tools Used 

in this Analysis, for a description of MAGIC). 

The Northeast, the Adirondack Mountains, and the 

Southeast were chosen for this modeling assessment 

because they are among the most acid-sensitive 

ecosystems and are located downwind of many of 

the emission sources affected by Title IV (see Figure 

4-8). The best-available environmental data were 

used to calibrate MAGIC (e.g., water quality, soil, 

deposition). In addition, MAGIC was calibrated 

using data collected on acidification in lakes and 

streams by the National Surface Water Survey 

(NSWS) and other programs, such as EMAP and 

the TIME and LTM programs. Both the NSWS 

and EMAP used statistical methods to sample 

a representative number of lakes and streams 

characteristic of the water quality and condition 

for all lakes or streams in each study region. 

For example, the NSWS approach represents 

approximately 7,100 lakes in the Northeast 

(Kaufmann et al., 1988) and 73,500 stream km 

in the southeast United States (U.S. EPA, 1988). 

Therefore, within the limitations of these surveys 

(only lakes larger than 4 hectares are included), 

the MAGIC results are indicative of the level of 

acidification in lakes and streams throughout these 

regions. For the analysis included in this report 

the MAGIC results represent approximately 1450 

lakes in the Adirondack Mountains, 1200 lakes in 

the Northeast, and 47,500 km of streams in the 

Southeast. Although there are some uncertainties 

with regard to the model, particularly concerning 

watershed nitrogen dynamics, MAGIC provides a 

generally accurate, well-tested, and widely accepted 

tool for modeling the response of surface water 

chemistry to sulfur and nitrogen deposition.

Climate conditions and emissions under the 2020 

Base Case scenario and future control Scenarios A, 

B, and C were assumed to be held constant after 

2020. Because of time lags inherent in ecological 

response to changes in pollutant emissions and 

deposition, lake and stream water quality in 2020 

would represent only a small portion of the recovery 

expected as a result of emission reductions included 

in the scenarios analyzed. Therefore, lake and 

stream conditions are presented for the year 2050, 

Lake Tear of the Clouds in the Adirondack 

Mountains (photo courtesy of New York State 

Department of Environmental Conservation).
NAPAP Report to Congress: An Integrated Assessment

83

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Adirondack Mountains (130 Lakes)

Northeast (98 Lakes)

Southeast (143 Streams)

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Study Regions

Figure 4-8. Regions and waterbodies modeled using MAGIC (prepared by U.S. EPA).

allowing the emission reductions to take effect more 

fully. Regional forest ecosystem responses to changes 

in deposition were not modeled in this assessment. 

Changes in lake and stream water quality for 

the three reduction scenarios (A, B, and C) were 

modeled in this analysis. The implementation of 

all scenarios is expected to significantly reduce 

the remaining percentage of lakes and streams 

in the acute and elevated concern categories in 

all three areas beyond what would occur with 

implementation of current emission-reduction 

programs included in the Base Case scenario. 

Although the effects of emission reductions 

beyond the Base Case scenario differ by region, 

the amount of reductions results in a modest 

change in the number of waterbodies of acute 

concern in all regions. The greater the emission 

reductions, the larger the number of lakes 

and streams of elevated concern that improve 

and move into the moderate or low concern 

categories (see Figure 4-1 for an explanation 

of the stages of recovery from acidification).

Lakes in the Northeast and the Adirondack 

Mountains change relatively quickly in response 

to changes in deposition, resulting in shorter 

recovery times from acidification than streams in 
Chapter 4 | Beyond Title IV: Ecological Impacts of Further Emission Reductions

84

the Southeast. Modeling results indicate that the 

Base Case scenario would result in 6% and 8% 

of lakes remaining in the acute concern class in 

the Northeast and the Adirondack Mountains, 

respectively, by 2050 (Figure 4-9). Although these 

results would improve the health of fish populations 

and other acid-sensitive species, 40% and 14% of 

modeled lakes in the Adirondack Mountains and 

the Northeast, respectively, still would remain in 

the elevated concern class and experience episodic 

acidification (Figure 4-10) under the Base Case 

scenario. With implementation of Scenario A, water 

quality conditions improve in the Northeast and 

the Adirondack Mountains, but remain similar to 

those in the Base Case scenario. The percentage of 

lakes of acute concern would decrease to 5% for the 

Northeast and 4% for the Adirondack Mountains 

from the levels in the Base Case scenario, while 

11% and 37% of modeled lakes would remain of 

elevated concern in these regions, respectively. A 

near elimination of lakes of acute concern in the 

Northeast and the Adirondack Mountains by 2050 

only occurs given the stringent emission-reduction 

scenarios of Scenarios B and C. Under Scenarios 

B and C, the percentage of lakes in the elevated 

concern class would also decrease, which is expected 

to improve the health of fish populations and allow 

other acid-sensitive species to return. With the 

implementation of Scenario C, the proportion of 

the lakes of elevated concern would decrease from 

40% to 27% in the Adirondack Mountains and 

from 14% to 10% in the Northeast. Given that 

some lakes are naturally acidic because of organic 

acids, it is not expected that all lakes in a region will 

have an ANC level greater than 50 μeq/L. MAGIC 

also can estimate the water quality condition before 

anthropogenic acidification started, providing a 

target “natural” ANC level for assessing whether 

a population of lakes or streams has achieved full 

recovery. Under Scenario C, the percentage of lakes 

in the elevated concern class in 2050 is closer to 

modeled preacidification levels than under any other 

scenario, particularly for lakes in the Northeast. 

This suggests the scale of emission reductions under 

Scenario C would approach full protection of 

lakes in the Northeast and Adirondack Mountains 

from acid deposition (Figures 4-9 and 4-10). 

8%

6%

4%

5%

1%

3%

1%

3%

15%

12%

9% 9%

0%

2%

4%

6%

8%

10%

12%

14%

16%

Adirondack Mountains Northeast Southeast

Lakes and Streams of Acute Concern 

Base Case

Scenario A

Scenario B

Scenario C

Notes: The area of the Northeast modeled by MAGIC includes Maine, New Hampshire, Vermont, Massachusetts, Rhode Island, 

Connecticut, New York (not including the Adirondacks), and northeastern Pennsylvania. The area of the Southeast includes Virginia, 

West Virginia, western North Carolina, eastern Tennessee, northern Georgia, and northwestern South Carolina. 

The Base Case scenario includes emission reductions as a result of CAIR and other mobile source regulations nalized after 2005. 

Figure 4-9. Projected changes in the number of lakes and streams of acute concern (ANC < 0 μeq/L) in 2050 with 

additional SO2 and NOx emission reductions (prepared by U.S. EPA). 

The story is somewhat different for the Southeast 

(Figures 4-9 and 4-10). Because of the unique 

watershed characteristics of the area, southeastern 

streams manifest changes over a longer period of 
NAPAP Report to Congress: An Integrated Assessment

85

time in response to emission reductions. Sulfur 

retention in the clay-rich soils of the Southeast 

and the eventual release of sulfur back into the 

surface water is an important reason for the delayed

ecosystem response in this region. As a result, 

modeling indicates that 15% of streams would 

remain of acute concern in 2050 in response to 

emission levels under the Base Case scenario. 

With implementation of Scenario A, water quality 

conditions improve in the Southeast, with the 

percentage of streams of acute concern decreasing 

to 12% of modeled streams by 2050. Under 

reductions of 75% and 90% in SO2 emissions 

from power generation sources (i.e., Scenarios B 

and C) beyond the levels in the Base Case scenario, 

additional recovery is expected, and streams in 

the acute concern class drop to 9% of modeled 

streams by 2050. Under Scenarios A, B, and C, 

the percentage of streams in the elevated concern 

class also would begin to decrease, showing a 

move toward ecosystem recovery. However, even 

under the large emission reductions included 

in Scenario C, 13% of modeled streams would 

remain of elevated concern in 2050 (Figure 

4-10). This lag in the recovery of streams of acute 

concern in the Southeast is due primarily to the 

large amount of sulfur that has been adsorbed by 

southeastern soils from decades of acid deposition. 

This adsorbed sulfur is predicted to be slowly 

released into streams over time, slowing the rate 

of stream recovery. For this reason, the response of 

streams in the Southeast is expected to lag emission 

reductions to a greater extent than lakes in the 

Northeast and the Adirondack Mountains, and 

the emission reductions in Scenarios A, B, and C 

will take longer to yield results in this region.

Critical Loads

Critical loads were calculated using a long-term 

steady-state model for the EMAP lakes in the 

Adirondack Mountains of New York under the 

Base Case scenario and Scenarios A through C (see 

Adirondack Mountains Critical Load Case Study in 

Chapter 2 for more details on the lakes modeled 

here). The analysis focuses on the combined load 

of sulfur and nitrogen deposition to which a lake 

could be subject and still support a moderately 

healthy aquatic ecosystem (i.e., ANC greater than 

50 μeq/L). Some lakes have naturally low acidic 

19%

14%

40%

16%

11%

37%

16%

11%

35%

13%

10%

27%

0%

10%

20%

30%

40%

50%

60%

Adirondack Mountains Northeast Southeast

Lakes and Streams of Elevated Concern

Base Case

Scenario A

Scenario B

Scenario C

Notes: The area of the Northeast modeled by MAGIC includes Maine, New Hampshire, Vermont, Massachusetts, Rhode Island, 

Connecticut, New York (not including the Adirondacks), and northeastern Pennsylvania. The area of the Southeast includes Virginia, 

West Virginia, western North Carolina, eastern Tennessee, northern Georgia, and northwestern South Carolina. 

The Base Case scenario includes emission reductions as a result of CAIR and other mobile source regulations nalized after 2005. 

 Modeled preacidication level

Figure 4-10. Projected changes in the number of lakes and streams of elevated concern (ANC = 0–50 μeq/L) in 

2050 with additional SO2 and NOx emission reductions (prepared by U.S. EPA). 
Chapter 4 | Beyond Title IV: Ecological Impacts of Further Emission Reductions

86

conditions. The preacidification ANC levels of the 

lakes were estimated using the MAGIC model. 

Of the 117 EMAP lakes modeled, six lakes had 

preacidification ANC levels below an ANC of 50 

μeq/L. Because of their natural acidity levels, it is 

unlikely that these six lakes would reach an ANC 

of 50 μeq/L or greater. These lakes were removed 

from this critical load analysis. Overall, this analysis 

shows that the future emission reductions (i.e., 

Scenarios A through C) would result in significant 

ecological recovery, as defined by reaching an ANC 

level greater than 50 μeq/L, and would significantly 

increase ecosystem protection of lakes in the 

Adirondack Mountains from acidic deposition.

Figure 4-11 shows lakes where deposition exceeds, 

does not exceed, and would be within 10% of 

the critical load. For Adirondack Mountain 

lakes under the Base Case scenario, 13% of lakes 

received levels of combined sulfur and nitrogen 

deposition that exceeded the critical load and 

could not be neutralized by the environment. 

For Scenarios A through C, 6%, 5%, and 1%, 

respectively, of the lakes would continue to receive 

acid deposition levels that exceeded the critical 

load. These results indicate that additional emission 

reductions lead to further ecosystem protection 

of lakes in the Adirondack Mountain region.

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Base Case Scenario A

Scenario B Scenario C

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Critical Load

Exceedances

Does Not Exceed

Critical Load

Within 10% of the

Critical Load

Exceeds Critical •

Load

•

•

Note: Critical load exceedances for sulfur and nitrogen for 111 lakes in the Adirondack Mountains under Base Case scenario and Scenarios A through C.

Figure 4-11. Critical load exceedances in the Adirondack Mountains (prepared by U.S. EPA). 
87

SUMMARY

Title IV has been successful in reducing emissions of SO2 and NOx from power generation 

to the levels set by Congress. In fact, by 2009, SO2 emissions from power plants were already 

3.25 million tons lower than the final 2010 cap level of 8.95 million tons, and NOx emissions 

were 6.1 million tons less than the projected level in 2000 without the ARP, or more than triple 

the Title IV NOx emission-reduction objective. As a result of these emission reductions, air quality 

has improved, providing significant human health benefits, and acid deposition has decreased to 

the extent that some acid-sensitive areas are beginning to show signs of recovery. Current emission 

reductions and the passage of time, which is needed for affected ecosystems to respond to the new 

environmental conditions, are expected to allow more acid-sensitive areas to recover. However, 

current emission-reduction levels (rules finalized as of spring 2005) are not sufficient to allow 

full recovery of acid-sensitive ecosystems. Estimates from modeling presented in this report show 

that additional emission reductions are necessary in order to protect acid-sensitive ecosystems. 

This report analyzes a range of SO2 and NOx emission-reduction scenarios to evaluate the extent 

to which further reductions could achieve additional environmental recovery and minimize 

the adverse ecological effects associated with acid rain. The results of the modeling presented 

in this report indicate that broader recovery is not predicted without additional emission 

reductions. To fully protect acid-sensitive ecosystems in all regions affected by acid deposition 

will require reductions in acid-yielding emissions from source categories beyond power plant SO2

emissions. The information presented in this report provides part of what is needed to determine 

appropriate future action; other information that is needed includes the costs and other impacts 

of emission reductions, the role of climate change and multiple pollutant interaction, and the 

value the public places on further improvements to the environment and human health. 

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