EPA 620/R-03/001 

January 2003 

Response of Surface Water Chemistry to 

the Clean Air Act Amendments of 1990 

U.S. Environmental Protection Agency 

Office of Research and Development 

National Health and Environmental Effects Research Laboratory 

Research Triangle Park, NC 27711
ii 

Notice 

The research described in this report has been funded by the United States Environmental 

Protection Agency. This document has been reviewed according to Agency policy and approved 

for publication. Mention of trade names or commercial products does not constitute 

endorsement or recommendation for use. 
iii 

Dear Reader: 

 EPA’s Office of Research and Development and collaborators are issuing a report, 

“Response of Surface Water Chemistry to the Clean Air Act Amendments of 1990,” 

EPA/620/R-02/004. This report suggests that a market-based approach to pollution control 

works. The positive results are a product of the 1990 Clean Air Act Amendments (CAAA), 

which utilize the successful market-based cap and trade program on which President Bush’s 

Clear Skies Initiative is modeled. 

 The report concludes that the CAAA regulations have resulted in a large and widespread 

decrease in the deposition of wet sulfur. The amount of wet sulfur deposited in lakes and 

streams declined by approximately 40 percent in the 1990s for all of the regions. Regional 

declines in surface water sulfate can be directly linked to declines in emissions and deposition of 

sulfur that have occurred since the 1990 CAAA. 

 The 1990 CAAA required control measures for coal-fired power plants, industries, and 

other sources in an effort to reduce sulfur emissions that contribute to the development of acid 

rain. EPA and collaborators have conducted extensive monitoring and scientific assessment 

since 1990 to determine whether control measures have reduced levels of acidity in lakes and 

streams in five regions of the Northern and Eastern United States most affected by acid rain. 

 In three regions, one-quarter to one-third of lakes and streams previously affected by acid 

rain are no longer acidic. That’s good news for the environment and good news for the marketbased approach. 

Sincerely, 

Paul Gilman, Ph.D. 

Assistant Administrator 
iv 

Authors

John L. Stoddard1, Jeffrey S. Kahl2, Frank A. Deviney3, David R. DeWalle4, Charles T. 

Driscoll5

, Alan T. Herlihy6, James H. Kellogg7, Peter S. Murdoch8, James R. Webb9, and 

Katherine E. Webster10

1 Western Ecology Division 

U.S. Environmental Protection Agency 

200 SW 35th Street 

Corvallis, OR 97333 

stoddard.john@epa.gov 

2 Senator George Mitchell Center 

102 Norman Smith Hall 

University of Maine 

Orono, ME 04469 

kahl@maine.edu 

3 Department of Environmental Sciences 

Clark Hall 

University of Virginia 

Charlottesville, VA 22903

4 Pennsylvania State University 

106 Land and Water Research Bldg. 

University Park, PA 16802 

5 Dept. Civil & Environmental Engineering 

Syracuse University 

Syracuse, NY 13244-1190

6 Oregon State University 

c/o U.S. EPA 

200 SW 35th Street 

Corvallis, OR 97333 

7 Dept. Environmental Conservation 

Water Quality Division 

103 South Main Street 

Waterbury, VT 05671-0408 

8 U.S. Geological Survey 

425 Jordan Road 

Troy, NY 12180

9 Department of Environmental Sciences 

Clark Hall 

University of Virginia 

Charlottesville, VA 22903 

10 Department of Biological Sciences 

Murray Hall 

University of Maine 

Orono, ME 04469

Reviewers

Richard Haeuber, U.S. EPA Office of Air and Radiation 

Jacques Dupont, Quebec Ministry of the Environment 

Dean Jeffries, Environment Canada 

Kathy Tonnessen, National Park Service 

Gary Lovett, Institute of Ecosystem Studies, Millbrook, NY 

N. Scott Urquhart, Department of Statistics, Colorado State University 

Acknowledgements

The quality of this report was improved significantly by the comments of the reviewers. The 

authors greatly appreciate their efforts and constructive input. Supplemental funding from the 

U.S. EPA through a USGS interagency agreement (DW14938374) made it possible for co-author 

Kahl to participate in this effort. 
v 

Table of Contents 

Executive Summary...................................................................................................................... vii 

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

Objectives ........................................................................................................................................3 

Background......................................................................................................................................8 

Data Used in this Report................................................................................................................12 

Trends in Emissions and Deposition..............................................................................................20 

Response of Surface Waters, 1990-2000.......................................................................................29 

Conclusions....................................................................................................................................64 

Terms and Acronyms.....................................................................................................................68 

References......................................................................................................................................70 

List of Figures 

Figure A. Acid sensitive regions of the northern and eastern United States ............................... vii 

Figure B. Summary of regional trends in surface water chemistry .............................................. ix 

Figure 1. Emission sources affected under the EPA Acid Rain program .....................................3 

Figure 2. Acid sensitive regions of the northern and eastern United States .................................6 

Figure 3. Acid sensitive regions of the eastern United States........................................................7 

Figure 4. Distribution of fish species by lake pH ........................................................................11 

Figure 5. Trends in sulfur dioxide emissions following implementation of Phase I ...................21 

Figure 6. Trends in wet sulfate deposition in the eastern United States......................................21 

Figure 7. Trends in wet deposition concentrations of sulfate during 1990-2000 ........................23 

Figure 8. Trends in nitrogen oxide emissions following implementation of Phase I ..................24 

Figure 9. Trends in wet NO3

-

 deposition in the eastern United States.........................................25 

Figure 10. Cumulative frequency distribution of slopes for sulfate and nitrogen ........................26 

Figure 11. Distribution of slopes for acidity and base cations trends...........................................28 

Figure 12. Annual precipitation in the study regions....................................................................30 

Figure 13. The relationship between Gran ANC and [CA2+ + Mg2+]...........................................31 

Figure 14. Time series data...........................................................................................................34 

Figure 15. Time series data...........................................................................................................35 

Figure 16. Time series data...........................................................................................................36 

Figure 17. Time series data...........................................................................................................37 

Figure 18. Time series data...........................................................................................................38 

Figure 19. Regional sulfate trends in LTM network.....................................................................40 

Figure 20. Regional nitrate trends in LTM network.....................................................................41 

Figure 21. Regional ANC trends in LTM network.......................................................................44 

Figure 22. Regional hydrogen ion trends in LTM network..........................................................45 

Figure 23. Regional [CA2+ + Mg2+] trends in LTM network........................................................47 

Figure 24. Regional DOC trends in LTM network.......................................................................48 

Figure 25. Relationship between trends in DOC and trends in SO4

2-...........................................49 

Figure 26. Regional aluminum trends in LTM network...............................................................50 
vi 

Figure 27. Comparison of trends in SO4

2- in TIME probability sites ...........................................54 

Figure 28. Comparison of trends in ANC in TIME probability sites ...........................................55 

Figure 29. Comparison of trends in base cations in TIME probability sites ................................57 

Figure 30. Comparison of % change in SO4

2- concentration........................................................58 

Figure 31. Relationship between summer and spring ANC values at LTM sites.........................61 

Figure 32. Paleolimnological reconstruction of historical pH change..........................................63 

Figure 33. The interaction of SO4

2- , NO3- and base cations in East Bear Brook.........................65 

Figure 34. Regional trends, 1990-2000.........................................................................................66 

List of Tables 

Table 1. Sources of data and sample sizes for datasets..................................................................5 

Table 2. Location of NADP/NTN wet deposition sites ...............................................................14 

Table 3. Typical analytical methods used for samples ................................................................19 

Table 4. Regional trend results for atmospheric deposition ........................................................27 

Table 5. Regional trend results for Long-Term Monitoring sites................................................39 

Table 6. Slope of trends in Gran ANC.........................................................................................52 

Table 7. Regional trend results for populations of sites in acid sensitive regions.......................53 

Table 8. Estimates of change in number and proportion of acidic surface waters ......................62 
vii 

EXECUTIVE SUMMARY 

Response of Surface Water Chemistry to the Clean Air Act Amendments of 1990 

Purpose of this report. Title IV of the 1990 Clean Air Act Amendments (CAAA) set target 

reductions for sulfur and nitrogen emissions from industrial sources as a means of reducing the 

acidity in deposition. One of the intended effects of the reductions was to decrease the acidity of 

low alkalinity waters and thereby improve their biological condition. The purpose of this report 

is to assess recent changes in surface water chemistry in the northern and eastern U.S., in 

response to changes in deposition. The regions covered in this report are New England (sites in 

Maine, New Hampshire, Vermont and Massachusetts), the Adirondack Mountains of New York, 

the Northern Appalachian Plateau (New York, Pennsylvania and West Virginia), the Ridge and 

Blue Ridge provinces of Virginia, and the Upper Midwest (Wisconsin and Michigan). The data 

covered in this report are from 1990 through 2000, the period since the last major science review 

by the National Acidic Precipitation Assessment Program (NAPAP). 

Figure A. Acid sensitive regions of the northern and eastern United States; this report assesses 

trends in surface waters in each of these regions. 
viii 

Substantial reductions in emissions of sulfur have occurred in the past 30 years, with the rate of 

decline accelerated by Phase I of the 1990 CAAA, implemented in 1995. Modest reductions in 

nitrogen emissions have occurred since 1996. The key questions are (a) whether the declines in 

emissions translate into reductions in acidic deposition; and (b) whether biologically relevant 

water chemistry has improved in acid sensitive regions. The measures of expected recovery 

include decreased acidity, sulfate, and toxic dissolved aluminum concentrations. 

Anthropogenic acidity in atmospheric deposition. NOx and SOx from the combustion of fossil 

fuels react with water in the atmosphere to produce acid rain, a dilute solution of nitric and 

sulfuric acids. This acidity (and the acid anions sulfate and nitrate) may travel hundreds of miles 

before being deposited on the landscape. The northern and eastern U.S. receives precipitation 

with mean pH that ranges from 4.3 in Pennsylvania and New York, to 4.8 in Maine and the 

Upper Midwest. The acidity (hydrogen ion concentration) in precipitation in the eastern U.S. is 

at least twice as high as in pre-industrial times. Atmospheric deposition is one of the most 

ubiquitous non-point sources of chemicals to ecosystems. 

Acid-base status of surface waters. The 1984-86 EPA National Surface Water Survey (NSWS) 

estimated the number of acidic waters at 4.2% of lakes and 2.7% of stream segments in acidsensitive regions of the North and East. Acidic waters are defined as having acid neutralizing 

capacity (ANC) less than zero (i.e., no acid buffering capacity in the water), corresponding to a 

pH of about 5.2. 

This report addresses the recent chemical responses in the surface waters in five regions of the 

North and East that are considered sensitive to acidic deposition. The data in this report are 

largely from the EPA Long Term Monitoring (LTM) and the EPA Temporally Integrated 

Monitoring of Ecosystems (TIME) projects, part of EMAP (Environmental Monitoring and 

Assessment Program). The regions include lakes in the Adirondacks, central and northern New 

England, and the upper Midwest. Sensitive regions with small streams are found in the midAtlantic region, including the northern and central Appalachian Plateau and the Ridge and Blue 

Ridge provinces. Surface waters in most other regions are not sensitive to the impacts of 

acidification due to the nature of the local geology. 

Recent changes in atmospheric deposition. We evaluated the changes in atmospheric 

deposition from the five regions during 1990-2000, using National Atmospheric Deposition 

Program (NADP) data. Sulfate declined significantly at a rate between -0.75 and -1.5 

μeq/L/year. There was a sharp drop in sulfate concentrations in 1995 and 1996, followed by a 

modest increase in 1997-2000, in parallel with emissions. Nitrogen (nitrate + ammonium) 

declined slightly in the Northeast, and increased slightly in the Upper Midwest; most of these 

changes can be attributed to changes in nitrate deposition. Base cations in deposition, which are 

important for the neutralization of acidity in precipitation and in watersheds, showed no 

significant changes during the decade in the East, and increased slightly in the Upper Midwest. 

These changes in deposition are a continuation of trends that pre-date the 1995 implementation 

of Phase I of the CAAA, and are consistent with other recent published analyses of changes in 

regional deposition patterns. 

Recent changes in acid base status in surface waters. All regions except the Ridge/Blue Ridge 

province in the mid-Atlantic showed significant declines in sulfate concentrations in surface 
ix 

waters, with rates ranging from -1.5 to -3 μeq/L/year (Figure B). These declines were consistent 

with the decline in sulfate in precipitation. Nitrate concentrations decreased in two regions with 

the highest ambient nitrate concentration (Adirondacks, Northern Appalachian Plateau) but were 

relatively unchanged in regions with low concentrations. Dissolved Organic Carbon (DOC) 

increased in each region, potentially contributing natural organic acidity to offset the recovery 

from decreased acidity and sulfate in deposition. 

Acid neutralizing capacity is a key indicator of recovery, as it reflects the capacity of watersheds 

to buffer inputs of acidity. We expect increasing values of either ANC or pH (or both) in 

response to decreasing deposition of sulfur and nitrogen from the atmosphere. ANC increased in 

three of the regions (Adirondacks, Northern Appalachian Plateau and Upper Midwest) at a rate 

of +1 μeq/L/year, despite a decline in base cations (calcium + magnesium) in each region (Figure 

B). The decline in base cations offsets some of the decline in sulfate, and thus limits the increase 

in ANC or pH. In the Adirondacks and Northern Appalachians, surface water ANC and pH both 

increased significantly in the 1990s; toxic aluminum concentrations also declined slightly in the 

Adirondacks. Regional surface water ANC did not change significantly in New England or in 

the Ridge/Blue Ridge. 

Figure B. Summary of regional trends in surface water chemistry in regions covered by this report.
x 

Has the number of acidic waters changed? Modest increases in ANC have reduced the 

number of acidic lakes and stream segments in some regions. We estimate that there are 

currently 150 Adirondack lakes with ANC less than 0, or 8.1% of the population, compared to 

13% (240 lakes) in the early 1990s. In the Upper Midwest, an estimated 80 of 250 lakes that 

were acidic in mid-1980s are no longer acidic. TIME surveys of streams in the northern 

Appalachian Plateau region estimated that 5,014 kilometers of streams (ca. 12%) were acidic in 

1993-94. We estimate that 3,600 kilometers of streams, or 8.5%, remain acidic in this region at 

the present time. In these three regions, approximately one-quarter to one-third of formerly 

acidic surface waters are no longer acidic, although still with very low ANC. We find little 

evidence of a regional change in the acidity status of New England or the Ridge/Blue Ridge 

regions and infer that the numbers of acidic waters remain relatively unchanged. There is no 

evidence that the number of acidic waters have increased in any region, despite a general decline 

in base cations and a possible increase in natural organic acidity. 

Do changes in deposition translate into changes in surface waters? A major goal of this 

assessment is to evaluate the effectiveness of emission reductions in changing surface water 

chemistry. We only make this assessment for sulfate because changes in the deposition of 

nitrogen have been minor. In New England, the Adirondacks and the Northern Appalachians, 

the percent declines in sulfate concentrations in precipitation were generally steeper than in 

surface waters. This is largely as expected and suggests that for a majority of aquatic systems, 

sulfate recovery exhibits a somewhat lagged response. However, the lakes and streams with the 

steepest declines in sulfate had very similar rates to those in deposition, indicating that the most 

responsive watersheds responded directly and rapidly to the sulfate decrease in deposition. As 

expected, there was little correspondence between rates of sulfate decline in streams and 

deposition in the Ridge and Blue Ridge provinces due to the adsorptive capacity of the soils in 

the region. In the upper Midwest, the rate of decline in lakes was greater than the decline in 

deposition, probably reflecting the residual effects of the drought of the late 1980s. Longer term, 

we expect the chemistry of seepage lakes in the Upper Midwest to mirror the decline in 

deposition, similar to the pattern seen in seepage lakes in New England that did not experience 

the 1980s drought. 

Complications for assessing recovery. Declines in atmospheric deposition of sulfate have led 

to nearly universal declines in sulfate concentrations in surface waters. This response is one 

simple measure of the intended recovery in surface waters and marks a success of the CAAA and 

efforts by industry in reducing SO2 emissions. However, the anticipated decrease in acidity 

corresponding to the decline in sulfate has been modest. 

It is important to recognize that recovery will not be a linear process. Moreover, the changes in 

surface water chemistry reported here have occurred over very short periods relative to the 

implementation of the CAAA emission reductions in 1995. The decline in sulfate is without 

question due to the decline in emissions and deposition, but mechanisms producing other 

changes are much less clear. Other responses in surface waters may be partially attributable to 

factors other than atmospheric deposition, such as climate change and forest maturation. In 

particular, some of the observed increase in ANC may result from decreases in nitrate 

concentrations (e.g., in the Adirondacks and Northern Appalachian Plateau); changes in nitrate 

are unrelated to changes in nitrogen deposition and are not expected to continue. If the trend 

toward lower nitrate in surface water reverses, some of the gains in ANC may be lost. 
xi 

We can identify at least five factors that are important in determining the recovery, or lack of 

recovery, in surface waters of the northern and eastern U.S. Continued long-term research and 

monitoring will be necessary to understand the causes, effects, and trends in these processes. 

1) Base cations. We report declining surface water concentrations of base cations (e.g., 

calcium, magnesium) in all of the glaciated regions in this report (the Ridge and Blue 

Ridge region is the only non-glaciated region). At some individual sites, further 

acidification has occurred because base cations are declining more steeply than sulfate. 

While decreases in base cation loss from watersheds probably indicates slower rates of 

soil acidification, they nonetheless limit the magnitude of surface water recovery. 

Continued long-term research at acid-sensitive sites is needed to determine the cause and 

effect of the relationship between base cations and sulfate and the effects of cation loss 

on soil and surface water recovery. 

2) Nitrogen. Continued atmospheric loading of nitrogen may be influencing the acid-base 

status of watersheds in yet undetermined ways. Unlike sulfate, concentrations of 

nitrogen in deposition have not changed substantially in 20 years. Also unlike sulfate, 

most nitrogen deposited from the atmosphere is retained in watershed soils and 

vegetation; nitrogen sequestration is not expected to continue ad infinitum (Stoddard 

1994, Aber et al., in press). We report that surface water nitrate concentrations are 

largely unchanged, except in two regions characterized by high nitrate concentrations a 

decade ago (Adirondacks, Northern Appalachian Plateau). The mechanisms behind 

these decreases in nitrate are not understood and could include climate change, forest 

recovery from disturbance, and the effects of land-use history. Future increases in 

nitrate concentrations in all regions are not improbable and would retard recovery if 

other factors remain constant. 

3) Natural organic acidity. Increases in dissolved organic carbon in acid-sensitive waters 

may have contributed additional natural organic acidity to surface waters, complicating 

our interpretation of the response in acidity. This factor is an important long-term 

research question that is probably linked to complex issues including climate change and 

forest maturation. 

4) Climate. Climatic fluctuations induce variability in surface water chemistry and thus 

obscure changes that we expect to result from declining acidic deposition. Climate or 

climate-related processes may counteract recovery by producing declines in base cations 

to offset a decline in sulfate or by inducing an increase in natural organic acidity. These 

interactions of factors underscore the need to continue monitoring a subset of sensitive 

systems so as to understand the full suite of drivers and responses in ecosystems. 

5) Lag in response. Documentation of the response of watersheds to changes in atmospheric 

deposition may take longer than the timeframe of available data. Recovery itself may 

have an inherent lag time beyond the time scale of currently available monitoring data. 

Moreover, the changes observed are not unidirectional. Uncertainty with respect to 

timeframes can only be resolved with continued long-term data. 

Indicators of recovery. A main goal of the Title IV of the CAAA is to decrease the acidity of 

affected surface waters. Although decreases in acidity have occurred in several regions, 

additional factors appear to point toward recovery, forecasting an improvement in biologically 
xii 

relevant surface water chemistry. It is not yet clear if further reductions in emissions and 

deposition will be necessary for widespread recovery to occur. These factors forecast the onset 

of recovery: 

a) Sulfate is an increasingly smaller percentage of total ion concentration in surface waters. 

b) ANC has increased modestly in three of the five regions. 

c) Dissolved Organic Carbon has increased regionally, perhaps toward a more natural preindustrial concentration as acidity decreases in surface waters. 

d) Toxic aluminum concentrations appear to have decreased slightly in some sensitive systems. 

Expectations for recovery. An important consideration for measuring the success of the CAAA 

is to have appropriate expectations for the magnitude of potential recovery. Lakes inferred to 

have been measurably acidified by atmospheric deposition were already marginally acidic, 

typically with pH less than 6, before anthropogenic atmospheric pollution began more than 100 

years ago. Therefore, full recovery of acidic lakes will not yield neutral pH. However, there is 

evidence that DOC will increase during recovery, and both increasing DOC and increasing pH 

values will lower the toxicity of aluminum. This change may allow recovery of fish populations 

to historical conditions even if pH remains low. 

Recommendations. In the North and East, there is evidence of recovery from the effects of 

acidic deposition. The complexities of ecosystem response – effects of forest health, soil status, 

natural organic acidity, the relative importance of sulfur vs. nitrogen deposition, future 

emission/deposition scenarios – make predictions of the magnitude and timing of further 

recovery uncertain. The results of this trend analysis suggest two recommendations for 

environmental monitoring: 

1) Deposition monitoring: The analyses in this report depended heavily on the long-term 

NADP/NTN program for monitoring the chemistry of precipitation. The future 

assessment of deposition and aquatic trends will depend heavily on these data, and 

therefore our recommendation is to maintain a national precipitation chemistry network. 

2) Surface water monitoring: The effectiveness of current or future amendments to the 

Clean Air Act can best be determined by monitoring the response of subpopulations of 

sensitive surface waters through time. Long-term records provide the benchmark for 

understanding trends in ecological responses. The reviewers of early drafts of this report 

strongly urged the authors to recommend the continuation of the long-term research 

programs upon which this report is based and the addition of biological monitoring to 

begin documenting potential biotic recovery. 

Future research. The data from these long-term sites will be invaluable for the evaluation of 

the response of forested watersheds and surface waters to a host of research and regulatory issues 

related to acidic deposition, including soil and surface water recovery, controls on nitrogen 

retention, mechanisms of base cation depletion, forest health, sinks for sulfur in watersheds, 

changes in DOC and speciation of aluminum, and various factors related to climate change. As 

one reviewer of this report noted, ”...these sites have irreplaceable long-term data that should 

constitute a ‘research infrastructure’ akin to an EPA laboratory. These sites will help address 

many basic science issues in which EPA ORD has a continuing interest....” Moreover, as 
xiii 

several of the reviewers observed, long-term data serve as the foundation for ecological research 

and modeling. Without such data, our ability to ask the right questions is reduced, and our 

ability to base the answers to these questions on actual data is likewise compromised.
1

INTRODUCTION 

The existence of ‘acid rain’, more properly 

called ‘acidic deposition’, has been known 

for more than 100 years (Smith 1872). 

Documentation of acidification of surface 

waters began in Scandinavia (e.g., Oden 

1968), although reports of acidic lakes date 

back to the 1950s in North America 

(Gorham, 1957). Recognition of the issue 

became common in the U.S. in the early 

1970s (Likens et al. 1972), with 

identification of impacts on fish by the mid1970s (e.g., Schofield 1976). 

Acidic deposition is a dilute solution of 

sulfuric acid and nitric acids derived from 

emissions of SO2 and NOx, by-products of 

the combustion of fossil fuels. The regions 

most impacted by acidic deposition are 

downwind of the documented sources of 

industrial sulfur emissions in the 

Midwestern states, and located in 

geologically-sensitive areas in the Northeast, 

Mid-Atlantic and Upper Midwest regions. 

Lake and stream chemical data from these 

regions are the subject of this report. 

The potential and documented effects of 

acidic deposition include damage to building 

materials, impacts to forests or crops, 

leaching of soil nutrients, acidification of 

surface waters, direct impacts on aquatic or 

amphibious biota, and indirect impact on 

other organisms in the same foodweb. The 

effects of the precursor emissions that lead 

to acidic deposition also include human 

respiratory stress and decreased visibility, 

both effects that would be reduced if S and 

N emissions were reduced. 

This report assesses the responses of surface 

waters in the five acid-sensitive regions of 

the northern and eastern states: New 

England (lakes); the Adirondack Mountains 

(lakes); the northern Appalachian Plateau 

(streams); the Ridge/Blue Ridge province 

(streams); and the Upper Midwest (lakes). 

With the exception of Florida, these regions 

include the vast majority of acidic surface 

waters in the U.S. Florida lakes are not 

addressed in this report because the majority 

of the acidity there is due to natural organic 

acidity (Baker et al. 1990), and because data 

have not been systematically collected there 

for more than a decade. 

There have been major changes in emissions 

during the past century. Emissions of SO2

are inferred to have increased from 9 million 

metric tons in 1900 to a peak of 29 million 

metric tons in 1973 (Husar et al. 1991), 

declining to 18 million metric tons by 1998. 

NOX emission are inferred to have increased 

from 2.5 million metric tons in 1900 to 22 

million metric tons around 1990, remaining 

relatively constant for more than a decade. 

As a result of the changes in emissions, the 

ratio of S to N in deposition in the North and 

East has changed from 2:1 to 1:1, with a 

minor decrease in acidity. It is within the 

context of these changes in deposition that 

this report evaluates the responses in surface 

waters since 1990. 

Aquatic Effects Research in the U.S. A 

consortium of federal agencies mobilized to 

address the issues of acidic deposition 

effects in 1980, creating the National 

Atmospheric Precipitation Assessment 

Program (NAPAP). The NAPAP effort 

spent more than $500 million in the 1980s, 

one of the major environmental research 

efforts in U.S. history. The ecological 

research under NAPAP greatly enhanced 

our understanding of ecological and 

hydrological processes relevant to 

acidification, and set the stage for the 

NAPAP assessment report, and the CAAA, 

in 1990. 
2

This report makes no attempt to summarize 

NAPAP research results from the 1980s 

because many such efforts exist (e.g. the 

State of Science and Technology series from 

1990, including reports 9 through 15 on 

surface water status, trends, and processes in 

watersheds). Instead, this report presents 

results as they relate to the patterns and 

trends found in the data in the 1990s. 

Title IV, Clean Air Act Amendments of 

1990. Despite the complication of longrange transport across political boundaries, 

legislative actions in the United State and 

Canada have resulted in reductions in acidic 

emissions and deposition. The purpose of 

Title IV (the Federal Acid Deposition 

Control Program) of the Clean Air Act 

Amendments of 1990 (Public Law 101-549, 

amending the Clean Air Act of 1970) were 

to reduce the adverse effects of acidic 

deposition through reductions in emissions 

of N and S acid pre-cursors (National Acid 

Precipitation Assessment Program 1998). 

The 1990 CAAA enhanced the trend toward 

decreased sulfur emissions and deposition 

that had been underway since the early 

1970s. 

Title IV largely targets the emissions of 

electric utilities that are estimated to account 

for 70% of S emissions and 30% of N 

emissions (Government Accounting Office 

2000). These reduction targets were 

implemented with Phase I in 1995, and 

Phase II in 2000 (Figure 1). The goal is a 10 

million ton reduction in SO2, and 2 million 

ton reduction in NOx, by 2010, compared to 

the baseline year of 1980. Emissions of SO2

from utilities were 10.6 million tons per year 

in 2001, compared to the baseline years of 

1980 (17.3 m tons) and 1990 (15.7 m tons). 

The target decrease in S emissions 

represents a 40% permanent cap compared 

to 1980 emissions. The target decrease in N 

is 10% compared to 1980. 

Phase I affected 263 units at 110 mostly 

coal-burning electric utility plants located in 

21 eastern and mid-western states. An 

additional 182 units joined Phase I of the 

program as substitution or compensating 

units, bringing the total of Phase I units to 

445. Phase II tightened the annual 

emissions limits imposed on these large, 

higher-emitting plants and also set 

restrictions on smaller, cleaner plants fired 

by coal, oil, and gas, encompassing over 

2,000 units in all. In 2001, program sources 

emitted 10.6 million tons of SO2 — nearly 7 

million tons below 1980 levels, representing 

a reduction of approximately 40%. 

Title IV focuses on NOx-emitting coal-fired 

electric utility boilers. As with the SO2

emission reduction requirements, the NOx 

program was implemented in two phases, 

beginning in 1996 and 2000. The NOx 

program embodies many of the same 

principles of the SO2 trading program, with 

flexibility in the method to achieve emission 

reductions. However, it does not cap NOx 

emissions as the SO2 program does, nor does 

it utilize an allowance trading system. There 

has been a reduction in NOx emissions from 

Title IV affected sources from 5.5 million 

tons in 1990 to 4.7 million tons in 2001. 

Linkages to other environmental 

stressors. Although the evaluation of the 

effectiveness of the Clean Air Act 

Amendments is important for understanding 

the direct link to surface water chemistry 

and fisheries, there are ancillary advantages 

to controlling emissions and deposition. For 

example, coastal eutrophication is a problem 

of increasing importance for estuarine 

ecosystems and marine productivity 

(Jaworski et al. 1997, Paerl 2002, Nielsen 

and Kahl in press).
3

Figure 1. Emission sources affected under the EPA Acid Rain program. 

Similarly, the linkage to climate change is 

impossible to ignore. Reductions in S and N 

emissions may be coupled to reduced 

emissions of greenhouse gases and mercury, 

especially if the reductions are the result of 

increased efficiency or conservation. In 

addition, potential climate change may have 

an impact on the response of ecosystems and 

complicate interpretation of data collected 

for purposes of evaluating the effectiveness 

of the CAAA (Webster and Brezonik 1995, 

Norton et al. in press).

OBJECTIVES 

The overarching objective of this report is to 

evaluate the chemical response of lakes and 

streams in the northern and eastern U.S. 

during the past decade. This response will 

be evaluated relative to the well-documented 

status and trends in precipitation chemistry. 

For purposes of this report, we define the 

northern and eastern U.S. as including all of 

the acid sensitive regions north of the 

Virginia/North Carolina border and east of 

the Rocky Mountains. The time period of 

our focus is 1990 to 2000. 
4

Our specific objectives are to: 

1) identify trends and regional patterns in 

surface water chemistry using: 

a. statistical population assessments 

from the U.S. EPA 

Environmental Monitoring and 

Assessment Program (EMAP), 

and the Temporally Intensive 

Monitoring of Ecosystems of 

Ecosystems (TIME) project – 

these data result from repeated 

probability surveys of lakes in 

the Adirondacks and New 

England and of streams in the 

Mid-Atlantic, and can be 

expanded statistically to the 

population(s) of surface waters 

in each region; 

b. data from lakes and streams 

identified to be the indicator 

systems most responsive to 

changes in deposition chemistry, 

including the U.S. EPA LongTerm Monitoring (LTM) project 

– these data are from specially 

selected sensitive surface waters, 

and represent the behavior of the 

most affected systems. 

2) evaluate linkages in changes in surface 

waters, if any, to changes in deposition 

that are related to regulatory goals; 

These two objectives will assist in providing 

the foundation for setting targets for future 

emission reductions, based on goals and 

expectations for future changes in aquatic 

chemistry. The results of analyses reported 

here will inform debates about reauthorization of the current CAAA, and 

about any future emissions regulation. 

Evaluating the response: a multi-aquatic 

system approach. A key issue for the 

evaluation of the CAAA is the relationship 

between trends in deposition and trends in 

surface water chemistry. Have regulatory 

changes produced the intended results in 

surface waters? The data may indicate a 

direct relationship between variables, a 

delayed response, or no apparent 

relationship. The more direct the apparent 

relationship, and the more types of aquatic 

systems in which the relationship occurs, the 

more confidence we have that the trends in 

deposition are the causal factor. 

The data in this report represent many 

locations and types of surface waters in the 

North and East (Table 1, Figures 2, 3); we 

group these sites into regions based on 

similarities in their acid sensitivity, and the 

rates of acidic deposition to which they have 

historically been exposed. Consistent 

recovery across a regional population of 

streams or lakes provides the strongest 

evidence that trends are directly related to 

acidic deposition controls. The regions 

covered here contained 95% of the acidic 

lakes, and 84% of the acidic streams, in the 

1980s (Linthurst et al. 1986a, Landers et al. 

1988). We are therefore confident that we 

have both represented the bulk of aquatic 

ecosystems affected by acidic deposition, 

and that we report regionally significant 

results. 

Acidic or acidified? In an assessment of 

recovery, it is important to distinguish 

between acidic waters and acidified waters. 

‘Acidic’ describes a condition that can be 

defined and documented, (i.e., Gran ANC ≤

0), and may be due either to the effects of 

acidic deposition, or to natural causes such 

as organic acidity or the weathering of 

sulfur-containing minerals in the watershed. 

In contrast, ‘acidified’ refers to the process 

of acidification (an increase in acidity 

observed through time), and does not require 

that the water body be acidic. The process 

of acidification has been observed in many 

long-term monitoring records (e.g., Lake 

Langtjern in Norway; Henriksen and Grande 

2002), in short-term studies of episodic 

acidification (e.g., Wigington et al. 1996), 
5

and experimental manipulations of aquatic 

ecosystems (Schindler 1988, Kahl et al. 

1993a, Wright et al. 1993). Chronic 

acidification can also be inferred from 

reconstructions of pH based on the remains 

of diatoms or chrysophytes collected in lake 

sediment cores (Charles et al. 1990, Smol et 

al. 1998). The term ‘anthropogenicallyacidified’ implies that acidic deposition was 

responsible for any increase in acidity; the 

term ‘acidified’ does not imply the cause.

Table 1. Sources of data and sample sizes for datasets used in this report, along with 

estimates of the condition of surface waters in each region in the 1980s. Statistical survey 

data are from the EMAP and TIME projects. Sensitive surface water data are from the LTM 

project, as well as other contributed studies. 

Sources of Data No. of sites1 Size of Population2 Percent acidic in 

1980s3

Statistical Surveys 

New England Lakes4 30 4,327 lakes 5% 

Adirondack Lakes4 43 1,290 lakes 14% 

Appalachian Plateau Streams 31 72,000 stream miles 6% 

Sensitive Surface Waters 

New England Lakes 24 N.A. 5% 

Adirondack Lakes 48 N.A. 14% 

Northern Appalachian Streams 9 N.A. 6% 

Upper Midwest Lakes 38 N.A. 3% 

Ridge/Blue Ridge Streams 69 N.A. 5% 

1 Number of monitoring sites with data available for this report (1990-2000); locations of sites are 

illustrated in Figures 2 and 3. 

2

 Total number of lakes, or stream length, for which statistical survey results can be inferred. Site 

selection for LTM (sensitive surface waters) is not statistically based, and results cannot be expanded to 

population level. 

3

 Estimates of extent of acidification, based on National Surface Water Survey results (Linthurst et al. 

1986b, Kaufmann et al. 1988). 

4 Estimates are for lakes with surface areas > 4 ha; estimates based on populations including smaller lakes 

are likely to be higher, due to the increased incidence of acidification in small lakes. 

What is ‘recovery’? Recovery from 

anthropogenic acidification can be defined 

in several ways. The most easily 

documented is a biologically-relevant 

decrease in acidity, measured as an increase 

in pH or Gran ANC, a measure of the 

buffering capacity of the water. However, 

an expectation of large increases in pH is 

unrealistic based on historical information 

for sensitive lakes. Today’s acidic lakes were 

marginally acidic in pre-industrial times 

(typical pH less than 6, inferred ANC 

probably less than 20 μeq/L; Charles and 

Norton 1986, Smol et al. 1998). Progress 

toward recovery should be considered to be 

progress toward pre-industrial chemical 

conditions, rather than progress toward pH 7.
6

Figure 2. Location of acid-sensitive regions of the northern and eastern U.S. covered in this report, and locations of in

sites used in trend analysis. 
7

Other indicators of recovery may be 

forecasting recovery, although it is not clear 

if further reductions in emissions and 

deposition will be necessary for widespread 

recovery to occur. These indicators of 

recovery include decreased sulfate 

concentrations, which constitute an 

increasingly smaller percentage of total ion 

concentration in surface waters, and 

decreased aluminum concentrations. Higher 

dissolved organic carbon (DOC) 

concentrations have been reported 

ubiquitously in surface waters of the 

northern hemisphere in the past decade 

(Evans and Monteith 2001, Skjelkvåle et al. 

2001, Jeffries et al. 2002; this report). 

Higher DOC may be an indicator of 

recovery, due to a decline in the protonation 

and precipitation of DOC from acidification 

(Krug and Frink 1983, Davis et al. 1985), or 

may be driven by climate change. These 

factors will be discussed later in the context 

of trends results for DOC and aluminum.

Figure 3. Location of acid-sensitive regions of the northern and eastern U.S. for which statistical survey 

data are available in the 1990s, and locations of individual TIME sites used in trend analysis. 
8

BACKGROUND 

Precipitation chemistry. ‘Acid rain’ has 

received much media coverage and research 

attention since widespread recognition in the 

late 1970s. ‘Acidic deposition’ refers to 

deposition of dilute acids from the 

atmosphere to the landscape. The source of 

the acids is the largely the combustion of 

fossil fuels that produce waste by-products 

including gases such as oxides of sulfur and 

nitrogen. Ammonia (NH3) is a by-product of 

some natural processes, as well as 

agricultural sources (e.g., application of 

nitrogen fertilizers; confined animal 

feedlots). In its dissolved form (NH4

+

) it 

contributes acidity to surface waters through 

the process of nitrification (Stoddard 1994). 

Oxidized sulfur and nitrogen gasses are acid 

precursors in the atmosphere. For example, 

SO2 reacts with water in the atmosphere to 

yield sulfuric acid: 

SO2 + H2O + 1⁄2O2 Æ H2SO4

An analogous reaction of water with 

nitrogen oxides, symbolized as NOx, yields 

nitric acid (HNO3). 

In addition to wet deposition (rain, snow, 

and fog), acidic deposition includes the 

deposition of dry, particulate, and gaseous 

acid precursors that become acidic in contact 

with moisture. This dry deposition is 

difficult to quantify and expensive to 

measure. Inferential methods indicate that 

dry deposition represents 20% to 80% of the 

total deposition of acids to the landscape, 

depending on factors such as location, 

season, and total rainfall (Sisterson et al. 

1990, Government Accounting Office 

2000). 

Natural sources can also contribute 

additional acidity to precipitation (Likens et 

al. 1983, Keene and Galloway 1984, 

Lindberg et al. 1984). Natural emissions of 

sulfuric, nitric, and hydrochloric acids occur 

from wetlands and geologic sources. Major 

natural sources of NOx include lightning 

and soil microbes. Organic acidity may 

arise from freshwater wetlands and coastal 

marshes. It is these natural sources that lead 

to the inference that pre-industrial 

precipitation in forested regions had a pH 

around 5.0 (Charlson and Rodhe 1982). If 

true, then modern precipitation in the North 

and East is two to three times more acidic 

than pre-industrial. 

The acidity of precipitation is still subject to 

misunderstanding. Even in pristine 

environments, precipitation pH is rarely 

controlled by the CO2 reaction that has an 

equilibrium pH of 5.6: 

H2O + CO2 Æ H2CO3

Because of the many sources of acidity in 

precipitation, pH 5.6 is not the benchmark 

‘normal’ pH against which the acidity of 

modern precipitation should be compared. 

Precipitation is a variable and complex 

mixture of particulates and solutes derived 

from local sources and long-range transport. 

For example, in arid or partly forested 

regions, dust from soil and bedrock typically 

neutralizes both the natural and human 

sources of acidity in precipitation, yielding a 

solution that may be quite basic (pH greater 

than 7). In the northeastern U.S. and eastern 

Canada, annual precipitation pH ranges from 

4.3 in Pennsylvania, New York, and Ohio, 

to 4.8 in Maine and maritime Canada 

(National Atmospheric Deposition 

Program/National Trends Network 2002). 

Effects of acidic deposition. Acidic 

deposition and emissions of acid precursors 

have been implicated as a risk to human 

health, including lung and cardiovascular 

system impacts (National Acid Precipitation 
9

Assessment Program 1998). These effects 

are typically associated with air pollution 

events and poor visibility caused by high 

ozone or sulfate haze. Acidic deposition is 

also known to cause degradation of building 

materials and monuments, such as those 

constructed from marble (National Acid 

Precipitation Assessment Program 1998). 

At the ecosystem-scale, impacts may occur 

to the terrestrial portions of both forested 

and agricultural landscapes. The most 

severe and widespread ecosystem effects are 

recognized to occur in surface waters, where 

both long-term chronic acidification and 

short-term temporary episodic acidifications 

are of concern. The issue of surface water 

acidification is the main subject of this 

report. 

Why do we care about lake and stream 

chemistry? Surface water chemistry is a 

direct indicator of the potential deleterious 

effects of acidification on biotic integrity. 

Because surface water chemistry integrates 

the sum of processes upstream in a 

watershed, it is also an indicator of the 

indirect effects of watershed-scale impacts, 

such as nitrogen saturation, forest decline, or 

soil acidification. Admittedly, we do not 

always know how to interpret the potentially 

complex interactions of upstream processes 

based solely on surface water chemistry, but 

their dual interpretive nature makes them 

invaluable in tracking surface water and 

watershed recovery. 

Chronic acidification. Surface waters 

become acidic when the supply of acids 

from atmospheric deposition and watershed 

processes exceeds the capacity of watershed 

soils and drainage waters to neutralize them 

(Charles 1991). Surface waters are defined 

as ‘acidic’ if their acid neutralizing capacity 

(ANC, analogous to alkalinity) is less than 

0, corresponding to pH values less than 

about 5.2. We are not addressing the 

impacts of acid-mine drainage in this report, 

and the sites used in this assessment exclude 

those with significant sources of geologic 

sulfur. The chemical conditions that define 

acidity are that acid anion concentrations 

(sulfate, nitrate, organic acids) are present in 

excess of concentrations of base cations 

(typically calcium or magnesium), the 

products of mineral weathering reactions 

that neutralize acidity in soil or rock. 

The National Surface Water Survey 

(NSWS) documented the status and extent 

of chronic acidification during probability 

surveys conducted from 1984 through 1988 

in acid sensitive regions throughout the U.S. 

(Linthurst et al. 1986b, Landers et al. 1987, 

Kaufmann et al. 1988). By using statistical 

techniques to select representative lakes and 

streams in each region surveyed, the NSWS 

estimated the chemical conditions of 28,300 

lakes and 56,000 stream reaches in all of the 

major acid-sensitive regions of the U.S. 

(Baker et al. 1990). The data collected by 

the NSWS have allowed us to focus our 

monitoring efforts on those areas that 

contain the vast majority of 

anthropogenically-acidified waters in the 

U.S., and that are most likely to respond to 

changes in deposition as a result of the 

CAAA. 

The NSWS concluded that 4.2% of lakes 

larger than 4 hectares and 2.7% of stream 

segments in the acid-sensitive regions were 

acidic. The regions represented in this 

report (Table 1) are estimated to contain 

95% of the lakes and 84% of the streams 

that have been anthropogenically acidified 

in the U.S. The Adirondacks had the largest 

proportion of acidic surface waters (14%) in 

the NSWS; the Adirondack Lake Survey 

Corporation sampled 1469 Adirondack lakes 

greater than 0.5 ha in size in 1984-87, and 

estimated that 26% were acidic (Driscoll et 

al. 1991). The proportions of lakes 

estimated by NSWS to be acidic were 

smaller in New England and the Upper 

Midwest (5% and 3%, respectively), but the 

large numbers of lakes in these regions 
10

translate to several hundred acidic waters in 

each region. 

The Valley and Ridge province and 

Northern Appalachian Plateau had 5% and 

6% acidic sites, respectively. The only acidsensitive region not assessed in the current 

report is Florida, where the high proportion 

of naturally acidic lakes, and a lack of longterm monitoring data, make assessment 

problematic. 

Episodic acidification. Short-term 

decreases in pH and Gran ANC, lasting on 

the order of hours to weeks, can occur 

during high flow events in both lakes and 

streams. Severe episodes can produce 

conditions that are as deleterious to biota as 

chronic acidification (Baker et al. 1996). 

The mechanisms that produce acidic 

episodes include dilution, and flushing of 

NO3

-

, SO4

2- and/or organic acids from forest 

soils (Kahl et al. 1992, Wigington et al. 

1996, Wigington 1999, Lawrence 2002). 

Acidic deposition may contribute to episodic 

acidification both by supplying nitrogen to 

sensitive watersheds (producing pulses of 

NO3

-

 during high flow events), producing 

hydrologically-mobile stores of SO4

2–

through dry deposition, and by lowering 

baseline pH and ANC, so that smaller 

episodes are sufficient to produce acidic 

conditions. 

Biologically-relevant Surface Water 

Chemistry. The main cause for concern 

over the effects of surface water 

acidification in the U.S. and elsewhere is the 

potential for detrimental biological affects 

(Baker and Christensen 1991, National Acid 

Precipitation Assessment Program 1998). 

Typically, there is concern for biological 

impact if the pH is less than 6. At low pH 

values, aluminum may be present at 

concentrations that are toxic to biota, 

including sensitive life stages of fish and 

sensitive invertebrates. Aluminum is an 

abundant and normally harmless component 

of rocks and soils. However, it leaches from 

silicate minerals when they come in contact 

with low pH waters. While much of the 

aluminum present in surface waters is 

organically-bound and relatively non-toxic, 

certain inorganic species are highly toxic. 

The best indicator of recovery in 

biologically-relevant chemistry would be a 

decrease in concentrations of inorganic 

monomeric aluminum, the most toxic form. 

Decreases in total aluminum would also 

suggest recovery, although the actual 

magnitude of the improvement in chemical 

conditions for biota would be unknown 

because we would not know how much of 

the decrease is due to inorganic vs. organic 

forms of aluminum. 

Biological effects. The biological effects of 

acidification have been demonstrated in 

laboratory and field bioassays (e.g., Baker et 

al. 1996), with whole-ecosystem 

acidification experiments (e.g., Schindler et 

al. 1985), and through field surveys (e.g., 

Baker and Schofield 1982, Gallagher and 

Baker 1990). A number of the species, 

especially of fish and macro-invertebrates, 

that commonly occur in surface waters 

susceptible to acidic deposition cannot 

survive, reproduce or compete in acidic 

waters. Sensitive species may be lost even at 

moderate levels of acidity. For example, 

some important zooplankton predators are 

not found at pH levels below 5.6; sensitive 

mayfly species (e.g., Baetis lapponicus) are 

affected at pH levels near 6.0; and sensitive 

fish species, such as the fathead minnow, 

experience recruitment failure and extinction 

at pH 5.6 to 5.9 (Baker and Christensen 

1991). 

Unfortunately, there are very few examples 

of long-term monitoring data for biological 

assemblages in acid-sensitive surface waters 

(e.g., Henriksen and Grande 2002), and none 

in the U.S. Therefore, many of our 

conclusions about the effect of acidic 

deposition on the distribution of sensitive 
11

species are based on relationships like the 

one shown in Figure 4. This relationship 

represents the sum total of the effects of 

acidification, lake size, and other complex 

interactions between the survival of fish 

species and the quality of their habitat. 

Research in the Adirondacks has shown that 

many lakes have always had marginal 

spawning habitat for Eastern Brook Trout 

(Schofield 1993), and at least some of the 

currently fishless acidic lakes in the 

Adirondacks may never have supported fish. 

Given the multiple lines of evidence 

available, it is very likely that acidic 

conditions, both natural and anthropogenic, 

limit the distribution of fish, benthic 

invertebrate, and zooplankton species, but a 

lack of historical data make it difficult to 

infer the magnitude of biological recovery 

(if any) from acidic deposition in individual 

lakes or streams. 

Ecological response: is sulfate or nitrate 

more deleterious? The NAPAP research 

effort of the 1980s was focused largely on 

sulfur as the problem behind the ecological 

impacts of acidic deposition (Baker et al. 

1990, Church et al. 1990, Turner et al. 

1990). Conceptual models were developed 

to understand regional responses (e.g., 

Galloway et al. 1983, Wright 1983). 

Computer models of watershed 

geochemistry enhanced our understanding of 

sulfate in watersheds (Christophersen and 

Wright 1981, Chen et al. 1984, Cosby et al. 

1985). As a result, sulfur biogeochemistry 

Figure 4. Distribution of fish species by lake pH in high elevation and seepage lakes in Maine 

(Kahl and Scott 1994). Y-axis represents the number of fish species found in each lake. 
12

is relatively well understood qualitatively. 

However, a decade after NAPAP, it is clear 

that we cannot predict sulfate quantitatively 

in most instances due to the complications 

of other, less well understood, factors 

One of the key factors still the subject of 

much research and scientific debate is the 

role of nitrogen in watershed responses to 

acidic deposition. The concept of nitrogen 

saturation (Aber et al. 1989, Stoddard 1994) 

received increasing attention in the 1990s 

(Murdoch and Stoddard 1992, Mitchell et al. 

1996, Williams et al. 1996, Aber et al. 

1998). Nitrogen saturation is defined as 

deposition of N to a watershed in excess of 

the assimilative capacity of soils and 

vegetation, resulting in the export of NO3

-

. 

Nitrate export can contribute to acidification 

(especially episodic acidification), 

mobilization of aluminum, and leaching of 

cations from soils (Aber et al. 1998, Fenn et 

al. 1998). 

In the U.S., nitrogen saturation has received 

the most attention in the eastern states, 

although it is not clear if chronic Nsaturation has occurred. Several forest 

manipulation studies in New England show 

increased mortality and declining growth of 

softwoods in response to N additions, 

leading to the suggestion that the widespread 

spruce decline in the eastern U.S. may be 

related to long-term N deposition (Aber et 

al. 1995). Nitrate concentrations in streams 

draining forests in the Northeast are higher 

than in any other forested region in the 

country, and have increased by a factor of 3 

to 4 since 1970 (Murdoch and Stoddard 

1992, Aber et al. in press), although these 

concentrations have begun to decrease based 

on analyses in this report. 

Our understanding of nitrogen cycling in 

forested watersheds is complicated by 

nitrogen’s response to climatic variation, 

land use history, forest disturbances (e.g., 

fire and insect defoliation) and drought 

(Mitchell et al. 1996, Aber and Driscoll 

1997, Goodale et al. 2000). Indeed, under 

most conditions, nitrogen from deposition 

functions as a fertilizer for N-limited forests. 

Despite the complexities of the nitrogen 

cycle in forested watersheds, the scientific 

consensus seems to be that, in the absence of 

substantial declines in nitrogen deposition, 

future increases in surface water NO3

–

concentrations, particularly in the Northeast 

and Mid-Atlantic U.S., are likely (Stoddard 

1994, Aber et al. in press). The difficulties 

inherent in controlling N emissions further 

complicate this issue—emissions are from 

both combustion sources (NOx) and 

agricultural sources (NH3 and NOx), and 

unlike S emissions, mobiles sources play a 

much larger role than stationary sources. An 

assessment of current status and trends, like 

the one presented here, can serve as the 

starting point for any future discussions of 

future effects and regulations concerning 

nitrogen. 

DATA USED IN THIS REPORT 

National Atmospheric Deposition 

Program/National Trends Program 

(NADP/NTN). This report relies heavily 

on long-term records of wet deposition from 

the NADP/NTN program 

(http://nadp.sws.uiuc.edu). The NADP/NTN 

network has collected data on rates of 

atmospheric deposition (rainfall, snowfall 

and major ion chemistry) since the early 

1980s. The network currently consists of 

245 sites, and is an excellent example of 

grass roots scientific cooperation creating an 

enormously useful source of data. 

In order to characterize deposition trends 

with adequate confidence, we chose to use at 

least 9 NADP/NTN locations for each 

region. None of the regional boundaries 

shown in Figure 2 enclose an adequate 

number of sites with long data records. We 

therefore expanded the regional boundaries 
13

slightly to include deposition monitoring 

stations within close enough proximity to 

have a high probability of exhibiting trends 

similar to those actually within the 

boundaries. The list of NADP/NTN sites 

used in this exercise are shown in Table 2. 

We chose not to use dry deposition 

estimates from programs such as the Clean 

Air Status and Trends Network (CASTNet; 

U.S. Environmental Protection Agency 

2002) for several reasons. Perhaps the most 

important of these is the need to compare 

directly the concentrations of ions in surface 

waters with the concentrations in deposition. 

At present there is no straightforward 

method to produce deposition concentrations 

(e.g., SO4

2– in μeq/L) from dry deposition 

data. In addition, the limited spatial and 

temporal coverage, and the high spatial 

variability of CASTNet data, makes them 

difficult to use with any confidence in the 

sort of regional analysis presented here. 

There are, for example, no CASTNet sites 

currently operating in the Adirondack 

Mountains, and only 2 useable sites in the 

Upper Midwest region, as we define it here. 
14

Table 2. Location of NADP/NTN wet deposition sites in each region. 

Region NADP/NTN 

Site Code 

State NADP/NTN Site Name 

New England MA01 Massachusetts North Atlantic Coastal Lab 

 MA08 Massachusetts Quabbin Reservoir 

 MA13 Massachusetts East 

 ME02 Maine Bridgton 

 ME09 Maine Greenville Station 

 ME98 Maine Acadia National Park 

 NH02 New Hampshire Hubbard Brook 

 VT01 Vermont Bennington 

 VT99 Vermont Underhill 

Adirondacks NY08 New York Auroroa Research Farm 

 NY20 New York Huntington Wildlife Refuge 

 NY52 New York Bennett Bridge 

 NY68 New York Biscuit Brook 

 NY98 New York Whiteface Mountain 

 NY99 New York West Point 

 VT01 Vermont Bennington 

 VT99 Vermont Underhill 

Northern Appalachians NY08 New York Auroroa Research Farm 

 NY65 New York Jasper 

 NY68 New York Biscuit Brook 

 NY99 New York West Point 

 PA15 Pennsylvania Penn State 

 PA29 Pennsylvania Kane Experimental Forest 

 PA42 Pennsylvania Leading Ridge 

 PA72 Pennsylvania Milford 

 WV18 West Virginia Parsons 

Upper Midwest MI09 Michigan Douglas Lake 

 MI98 Michigan Raco 

 MI99 Michigan Chassell 

 MN16 Minnesota Marcell Experimental Forest 

 MN18 Minnesota Fernberg 
15

Region NADP/NTN 

Site Code 

State NADP/NTN Site Name 

Upper Midwest MN23 Minnesota Camp Ripley 

 MN27 Minnesota Lamberton 

 WI09 Wisconsin Popple River 

 WI25 Wisconsin Suring 

 WI28 Wisconsin Lake Dubay 

 WI36 Wisconsin Trout Lake 

 WI37 Wisconsin Spooner 

 WI98 Wisconsin Wildcat Mountain 

 WI99 Wisconsin Lake Geneva 

Ridge and Blue Ridge KY22 Kentucky Lilley Cornett Woods 

 MD03 Maryland White Rock 

 NC45 North Carolina Mt. Mitchell 

 PA15 Pennsylvania Penn State 

 PA29 Pennsylvania Kane Experimental Forest 

 PA42 Pennsylvania Leading Ridge 

 PA72 Pennsylvania Milford 

 TN11 Tennessee Great Smoky Mountains National Park 

 VA00 Virginia Charlottesville 

 VA13 Virginia Horton's Station 

 VA28 Virginia Shenandoah National Park 

 WV04 West Virginia Babcock State Park 

 WV18 West Virginia Parsons 

Lake and stream monitoring programs. 

The core of the acid rain aquatic effects 

monitoring effort in the eastern U.S. has 

been two EPA programs: the Temporally 

Integrated Monitoring of Ecosystems 

(TIME) project (Stoddard 1990) and the 

Long-Term Monitoring (LTM) project (Ford 

et al. 1993, Stoddard et al. 1998b). Both 

projects are operated cooperatively with 

numerous collaborators in state agencies, 

academic institutions and other federal 

agencies. Augmenting these core programs 

are ongoing surveys of lake sub-populations 

especially sensitive to acidification or 

recovery. These programs all have slightly 

different objectives and structures, outlined 

below. 

TIME project: At the core of the TIME 

project is the concept of a probability 

sample, where each sampling site is chosen 

statistically to be representative of a target 

population. In the Northeast (New England 

and Adirondacks), this target population 

consists of lakes likely to be responsive to 
16

changes in rates of acidic deposition (i.e., 

those with Gran ANC < 100 μeq/L). In the 

Mid-Atlantic, the target population is upland 

streams with a high probability of 

responding to changes in acidic deposition 

(i.e., Northern Appalachian Plateau streams 

with Gran ANC < 100 μeq/L). 

Each lake or stream is sampled annually (in 

summer for lakes, in spring for streams), and 

results are extrapolated with known 

confidence to the target population(s) as a 

whole (Larsen and Urquhart 1993, Larsen et 

al. 1994, Stoddard et al. 1996, Urquhart et 

al. 1998). TIME sites were selected using 

the methods developed by EMAP (Paulsen 

et al. 1991, Herlihy et al. 2000). The TIME 

project began sampling Northeast lakes in 

1991; for the purposes of this report, we 

group these lakes into two acid-sensitive 

regions, the Adirondacks and New England. 

Data from 43 Adirondack lakes can be 

extrapolated to the target population of low 

ANC lakes in the region (there are ca. 1000 

low ANC Adirondack lakes, out of a total 

population of 1830 lakes with surface area > 

1 ha). Data from 30 lakes (representing ca. 

1500 low ANC lakes, out of a total 

population of 6,800) form the basis for our 

trend analysis in New England. Probabilistic 

monitoring of Mid-Atlantic streams began in 

1993; this report utilizes data from 30 low 

ANC streams in the Northern Appalachian 

Plateau (representing ca. 42,000 km of 

stream length, of which ca. 24,000 km are 

considered low ANC). In the Ridge and 

Blue Ridge provinces, the sample coverage 

from TIME was not sufficient to conduct 

regional trend analysis; all of our 

conclusions about this region (ca. 5,500 km 

of low ANC streams, out of a total 

population of 23,000 km) are therefore 

based on the non-probabilistic Virginia 

Trout Stream Survey described below. There 

are no sample survey data available from the 

Upper Midwest in the 1990-2000 time 

period. The ultimate goal of having a 

probability program like TIME is to 

determine not just how a representative 

sample of lakes and/or streams is changing 

through time, but to know whether the 

proportion of the population that is acidic 

has changed. 

LTM project: As a compliment to lake and 

stream sampling in a statistical population of 

lakes in TIME, the Long-Term Monitoring 

project samples a subset of sensitive lakes 

and streams with long-term data, most 

dating back to the early 1980s. These sites 

are sampled 3 to 15 times per year. This 

information is used to characterized how the 

most sensitive of aquatic systems in each 

region are responding to changing 

deposition, as well as giving information on 

seasonal chemistry and episodic 

acidification. In most regions, a small 

number of higher ANC (e.g., Gran ANC > 

100 μeq/L) sites are also sampled, and help 

separate temporal changes due to acidic 

deposition from those attributable to other 

disturbances (e.g., climate change, landuse 

change). Because of the long-term records at 

most LTM sites, their trends can also be 

placed in a better historical context than 

those of TIME sites, where data are only 

available from the 1990s. The details of data 

from each region are given below. 

New England lakes: The LTM project 

collects quarterly data from lakes in Maine 

(sampled by the University of Maine; Kahl 

et al. 1991, Kahl et al. 1993b) and Vermont 

(data collected by the Vermont Department 

of Environmental Conservation; Stoddard 

and Kellogg 1993, Stoddard et al. 1998a). 

Data from 24 New England lakes are 

available for the trend analysis reported here 

(time period 1990-2000). In addition to 

quarterly samples, a subset of these lakes 

have outlet samples collected on a weekly 

basis during the snowmelt season; these data 

are used to characterize spring episodic 

chemistry. The majority of New England 

LTM lakes have mean Gran ANC values 
17

ranging from –20 to 100 μeq/L; two higher 

ANC lakes (100< Gran ANC < 200 μeq/L) 

are also monitored. 

Adirondack lakes: Our trend analysis 

includes data from 48 Adirondack lakes, 

sampled monthly by the Adirondack Lake 

Survey Corporation (Driscoll and Van 

Dreason 1993, Driscoll et al. 1995); a subset 

of these lakes are sampled weekly during 

spring snowmelt to help characterize acidic 

episodes. The Adirondack LTM dataset 

includes both seepage and drainage lakes, 

most with Gran ANC values in the range –

50 to 100 μeq/L; 3 lakes with Gran ANC > 

100 μeq/L (but less than 200) are also 

monitored. 

Appalachian Plateau streams: Stream 

sampling in the Northern Appalachian 

Plateau is conducted roughly 15 times per 

year, with the samples spread evenly 

between baseflow (e.g., summer and fall) 

and high flow (e.g., spring) seasons. Data 

from 4 streams in the Catskill mountains 

(collected by the U.S. Geological Survey; 

Murdoch and Stoddard 1993), and 5 streams 

from Pennsylvania (collected by 

Pennsylvania State University; DeWalle and 

Swistock 1994) are used in this report. All 

of the Northern Appalachian LTM streams 

have mean Gran ANC values in the range 

-25 to 50 μeq/L. 

Upper Midwest lakes: Roughly 40 lakes in 

the Upper Midwest were originally included 

in the LTM project, but funding in this 

region was terminated in 1995. The 

Wisconsin Department of Natural Resources 

(funded by the Wisconsin Acid Deposition 

Research Council, the Wisconsin Utilities 

Association, the Electric Power Research 

Institute and the Wisconsin Department of 

Natural Resources) has continued limited 

sampling of a subset of these lakes, as well 

as carrying out additional sampling of an 

independent subset of seepage lakes in the 

state. The data used in this report consist of 

16 lakes (both drainage and seepage) 

sampled quarterly (Webster et al. 1993), and 

22 seepage lakes sampled annually in the 

1990s. All of the Upper Midwest data 

exhibit mean Gran ANC values from -30 to 

80 μeq/L. 

Ridge/Blue Ridge streams: Data from the 

Ridge and Blue Ridge provinces consist of a 

large number of streams sampled quarterly 

throughout the 1990s as part of the Virginia 

Trout Stream Survey (Webb et al. 1989, 

Webb et al. 1994), and a small number of 

streams sampled more intensively (as in the 

Northern Appalachian Plateau). A total of 

69 streams, all located in the Ridge section 

of the Ridge and Valley province, or within 

the Blue Ridge province, and all within the 

state of Virginia, had sufficient data for the 

trend analyses reported here. The data are 

collected cooperatively with the University 

of Virginia and the National Park Service. 

Mean Gran ANC values for the Ridge/Blue 

Ridge data range from –15 to 200 μeq/L, 

with 7 of the 69 sites exhibiting mean Gran 

ANC > 100 μeq/L. 

Supplemental datasets: Lake surveys 

conducted in the 1980s revealed that 70% of 

the known acidic lakes in the state of Maine 

were either in seepage lakes or high 

elevation lakes (Kahl et al. 1991). We have 

used data from these two sets of lakes, as 

well as from the control stream at the Bear 

Brook Watershed Manipulation Project 

(Norton et al. 1994, Church 1999), in some 

of the supplementary analyses presented in 

this report. 

Maine Seepage lakes: This dataset includes 

data from 120 of the estimated 150 lakes in 

Maine meeting the following criteria: 1) 

location in a sand and gravel mapped by the 

U.S.G.S. or Maine Geological Surveys; 2) 

depth at least 1 meter; and 3) area at least 

0.4 hectare (1 acre). Sampling was 

conducted in 1986-87 and 1998-2000, and 

included at least one fall index sample for 
18

each lake. The population was reduced to 

87 lakes for this report by screening for 

lakes with Gran ANC less than 100 μeq/L. 

Maine High Elevation lakes: Data from 90 

high elevation lakes (above 600 meters 

elevation), available annually for the periods 

1986-88 and 1997-99, represent the vast 

majority of lakes in Maine meeting this 

criteria: 1) depth at least 1 meter; and 2) area 

at least 0.4 hectare (1 acre). The population 

was reduced to 64 lakes for this report by 

screening for lakes with Gran ANC less than 

100 μeq/L. 

Surface water classification. For purposes 

of analysis and illustration, the surface water 

data contained in this report were classified 

according to average Gran ANC and DOC 

values. For the period of record, ‘acidic’

waters were those with mean Gran ANC less 

than or equal to zero. ‘Low ANC’ waters 

were those with Gran ANC greater than 0, 

but less than or equal to 25 μeq/L. 

‘Moderate ANC’ waters were those with 

Gran ANC greater than 25. The site 

selection process or the definition of the 

statistical population eliminated high ANC 

waters. The vast majority of sites included 

here all have mean Gran ANC less than 100 

μeq/L. 

Sites were also classified by dissolved 

organic carbon (DOC), with the intent to 

evaluate the contributions of natural organic 

acidity to the response of systems. Waters 

with mean DOC less than or equal to 5 mg/L 

were classified as ‘low DOC’. Sites with 

DOC greater than 5 mg/L were ‘high DOC’. 

The highest DOC lake (a seepage lake in 

Wisconsin) had a mean DOC concentration 

of 17 mg/L. 

Additionally, seepage lakes were also 

evaluated by silica concentration in order to 

classify systems as ‘perched’ or ‘flowthrough’. Seepage lakes with mean silica 

less than 1mg/L were considered perched. 

The significance of this classification is that 

perched lakes are disconnected 

hydrologically from groundwater (they 

supply water to the local groundwater) and 

have minimal watershed contributions to 

their chemistry. Such lakes are expected to 

be more responsive to changes in 

atmospheric deposition. Seepage lake data 

are available from Maine, the Adirondacks 

and Wisconsin. 

Analytical methods and quality 

assurance. The specific field protocols, 

laboratory methods, and quality assurance 

procedures are specific for each investigator. 

This information is contained in the cited 

publications of each research group and are 

compiled in Newell et al. (1987). The 

EMAP and TIME protocols and quality 

assurance methods are generally consistent 

with those of the LTM cooperators, and are 

detailed in Peck (1992) and in Table 3. 

The data used in this report have been 

screened for internal consistency among 

variables, including ion balance and 

conductance balance. Samples with 

unexplained variation in these variables 

have been deleted. Sites with mean Gran 

ANC greater than 200 μeq/L were deleted. 

We excluded sites with chloride values that 

were outliers in their region, because high 

Cl-

 is typically associated with human 

development in the watershed. The Cl- and 

associated Na+ would alter normal soil ion 

exchange relationships, thus obscuring the 

response to acidic deposition. 

Sea-salt corrections. We did not 

systematically correct surface water 

chemistry for contributions from marine 

aerosols, in large part because the marine 

influence was minor for most regions. The 

exception was stream data from the Bear 

Brook Watershed Manipulation, which is of 

close proximity to the Gulf of Maine, and 

where the detailed analyses of relationships 

between SO4

2– and base cation behavior 

required the greater precision provided by 
19

sea salt correction. For these streams, and in 

a few other instances as noted in the text, we 

corrected surface water chemistry by 

assuming all chloride was of marine origin, 

and correcting sulfate and base cations based 

on the ratios found in seawater (Riley and 

Chester 1971).

Table 3. Typical analytical methods used for samples reported here. Samples are filtered before analysis, 

except in Adirondack LTM. 

Variable Technique Protocol reference 

pH, closed cell Electrode Hillman, et al6, EPA 19.05

pH, aerated Electrode Hillman, et al6, EPA 5.05

Specific conductance Wheatstone bridge EPA 120.12,EPA 23.05

True color Spectrophotometer, 457.5 nm EPA 110.22

Gran ANC Gran Titration Hillman et al6, EPA 5.05

Anions: Cl-, NO3

-

, SO4

2- Ion chromatography EPA 300.01

Calcium AAS with N20-acetylene flame EPA 215.12

Magnesium AAS with N20-acetylene flame EPA 242.12

Sodium AAS with air-acetylene flame EPA 258.12

Potassium AAS with air-acetylene flame EPA 273.12

Aluminum, total AAS with graphite furnace EPA 200.94

Dissolved Organic Carbon IR C analyzer, persulfate oxidation EPA 415.12

Dissolved Inorganic Carbon IR C analyzer, direct injection EPA 13.05, OI, 19908

Ammonium Autoanalyzer EPA 9.05 and Bran & Luebbe 780-

86T7

Silica Autoanalyzer EPA 22.05 and Bran & Luebbe 785-

86T7

AAS=atomic absorption spectrophotometry 

IR=Infrared Spectrophotometry 

Method references: 

1

 Methods for the Determination of Inorganic Substances in Environmental Samples, EPA 600/R-93-100, 

1993. 

2

 Methods for Chemical Analysis of Water and Wastes, EPA 600/4-79-020, 1979, Revised 1983. 

3

 Standard Methods for Examination of Water and Wastewater, 18th ed. 1992. 

4

 Methods for the Determination of Metals in Environmental Samples, EPA 600/4-91/010, 1991, 

Supplement 1, EPA 600/R-94/111, 1994. 

5

 Handbook of Methods for Acid Deposition Studies: Laboratory Analysis For Surface Water Chemistry, 

EPA 600/4-87-026, 1987. 

6

 Hillman,D.C., J. Potter, and S. Simon, 1986. Analytical methods for the National Surface Water Survey, 

Eastern Lake Survey. EPA/600/4-86/009, EPA Las Vegas. 

7

 Bran & Luebbe Manual 

8

 OI Analytical Manual, 1990 
20

We did not correct NADP/NTN data for 

marine aerosols. Air masses move west to 

east, meaning that marine contributions to 

precipitation chemistry will be minor, with 

the exception of seasonal effects in coastal 

New England. Other sources are more likely 

to contribute chloride than seawater, and 

thus correction of sulfate and base cations 

based on a seawater-source assumption will 

likely lead to erroneous corrections. In the 

instance of base cation data that are 

important for watershed function, there is no 

point in mathematically removing cations 

presumed to arise from marine aerosols 

because the source of the cations is 

irrelevant to watersheds. 

Statistical analyses. Numerous statistical 

techniques are available to analyze trends in 

time series like those presented here. For 

trends at single monitoring sites, the 

Seasonal Kendall Tau (SKT) test (Hirsch et 

al. 1982, Hirsch and Slack 1984) is often the 

method of choice, because it deals well with 

censored data, and with data collected at 

irregular intervals with marked seasonality 

(Loftis and Taylor 1989). One weakness of 

the SKT is that does not compute a slope. 

The regional analyses we present in this 

report depend on the ability to calculate a 

robust estimator of slope for each site. 

Rather than utilizing a Sen estimator (Sen 

1968), as is often done with the SKT, we 

have chosen to use simple linear regression 

(SLR) to calculate a trend slope for each 

monitoring site. This has the added 

advantage of allowing us to use identical 

methods for detecting regional trends in both 

the LTM and TIME data; the probability 

data from TIME are not of sufficient sample 

size (for individual sites) or sampling 

frequency to make the choice of SKT 

appropriate. While the significance of 

individual tests conducted with SLR are 

questionable (especially when repeated tests 

are conducted within each region), the 

slopes calculated for multiple sites within a 

region represent a distribution of results, 

which can in turn be examined and analyzed 

for patterns. Within each region, we test for 

a significant regional trend by calculating 

confidence limits about the median value in 

the slope distribution (SAS Institute Inc. 

1988, Altman et al. 2000), and testing 

whether these confidence limits include 

zero. For a distribution in which all of the 

slopes are negative, for example, the median 

value would be significantly less than zero, 

indicating a significant regional downward 

trend. 

TRENDS IN EMISSIONS AND 

DEPOSITION 

Overview of Emissions and Deposition 

Changes: Sulfate emissions declined 

substantially in the 1990s (Figure 5); plants 

regulated under Phase I of the CAAA 

reduced emissions by 50%, while total 

emissions nationwide were reduced by 30% 

over 1980 levels.
21

Figure 5. Trends in sulfur dioxide emissions following implementation of Phase I of the Acid Rain 

Program: Total state-level utility SO2 (1980, 1990, 1999). 

Data from NADP/NTN indicate that sulfur 

emissions declines have contributed to a 

continuation of the long-term decline in 

sulfur deposition (Figure 6) from the 

previous decade (Holland et al. 1999, 

Shannon 1999, Government Accounting 

Office 2000, Lynch et al. 2000). Many 

NADP/NTN sites exhibit declines in excess 

of 45% since 1980. Compared to trends 

from 1983-94, the decline in wet deposition 

sulfate accelerated after 1995 at 74% of 

NADP sites in the eastern U.S. presumably 

as a result of Phase I of Title IV of the 

CAAA (Lynch et al. 2000).

 1989-1991 1995-1998 

Figure 6. Trends in wet sulfate deposition in the eastern United States from NADP/NTN data (1989-1991 

vs. 1995-1998). Data are in kg/ha. 
22

Power plant SO2 emissions increased after 

1996, perhaps because utilities overshot the 

CAAA targets in 1995 (Government 

Accounting Office 2000). This increase 

translated into a measurable increase in 

deposition in many NADP/NTN stations in 

the regions of interest for surface water 

response (Figure 7). The lowest SO4

2-

concentration in precipitation occurred in 

1995 or 1996, depending on the station 

(Figure 7; Government Accounting Office 

2000, National Atmospheric Deposition 

Program/National Trends Network 2002). 

Since 1980, total U.S. SO2 emissions have 

decreased by about 5.5 million tons. 

Decreases in NOx emissions were more 

modest than those of sulfur, with declines 

noted in several states in the northeastern 

corridor (Figure 6). Since 1990, total utility 

NOx emissions (Phase I and II sources) were 

reduced an average of 23% nationally, 

following implementation of Phase I of the 

Acid Rain Program. However, electric 

utilities contribute only about one-third of 

total NOx emissions. Since total NOx

emissions from other sources have remained 

relatively constant (motor vehicles and other 

industrial sources also contribute 

significantly), the reductions achieved under 

the Acid Rain Program have not resulted in 

a significant change in total NOx emissions.
23

 

Figure 7. Trends in wet deposition concentrations of SO4

2- during 1990 – 2000 (NADP, 2002). Boxes 

indicate the interquartile range in SO4

2- for each region in each year; caps indicate the 10th and 90th

percentiles; median concentration is indicated by line in box; dots enclose the range. 
24

Figure 8. Trends in nitrogen oxide emissions following implementation of Phase I of the Acid Rain 

Program: Total state-level utility NOx (1990 vs. 1999). 

Relatively constant NOx emissions translate 

to little change in wet NO3

–

 deposition in the 

1990s (Figure 8). The relative lack of 

change in NO3

–

 deposition, relative to the 

decline in SO4

2–

, has changed the ratio of 

S:N from 2:1 in 1980 to 1:1 in 2000 (Lynch 

et al. 2000). This long-term flat trend in 

nitrate deposition may have begun to shift 

downward after the implementation of Phase 

I for nitrate in 1996. Nevertheless, the 

geographic pattern of nitrate deposition has 

not changed substantially (Figure 9), in 

contrast to the changing pattern for sulfate 

(Figure 6). 

Regional deposition trends for SO4

2– and 

nitrogen: We evaluated the trends in wet 

deposition in each region for which we 

report trends in surface water response. Our 

analysis uses precipitation concentrations as 

reported by NADP/NTN (National 

Atmospheric Deposition Program/National 

Trends Network 2002). 

Throughout this report, we illustrate the 

range of trend behaviors in each region (for 

both deposition and surface waters) through 

the use of cumulative frequency 

distributions (e.g., Figure 10). These plots 

show the range of trend slopes in each 

region, with the minimum slope shown as 

the value of the x-axis corresponding to the 

left side of the regional curve, and maximum 

slope value equivalent to the x-value at the 

right end of the curve. The median value for 

each region, which is the value generally 

reported in the text, and for which statistical 

significance is calculated, can be estimated 

by finding the 50th percentile value for the 

region (from the y-axis) and estimating the 

corresponding value on the x-axis. For 

example, the median change in SO4

2–

concentration in wet deposition in the 

Northern Appalachian Plateau region 

(Figure 10a) is ca. -1.6 μeq/L/year.
25

Figure 9. Trends in wet NO3

-

 deposition in the eastern United States from NADP/NTN monitoring data 

(1989-1991 vs. 1995-1998). Data are in kg/ha. 

Sulfate concentrations declined substantially 

in each of the regions at a median rate 

between -0.8 and -1.5 μeq/L/year for the 

period 1990 to 2000 (Figure 10a). Nitrogen 

(NO3

-

 plus NH4

+

) declined slightly in the 

Northeast and Northern Appalachian 

Plateau, and was essentially unchanged in 

the Ridge and Blue Ridge provinces. The 

median change in wet deposition of nitrogen 

in the Upper Midwest during the 1990s was 

upward (Figure 10b); this was the only 

region to exhibit an increase in 

concentrations of either SO4

2– or nitrogen, 

consistent with results reported by Lynch et 

al. (2000). 

An immediate goal of the Title IV of the 

CAAA is to decrease atmospheric 

deposition of SO4

2- by reducing sulfur 

emissions. Wet deposition of SO4

2- declined 

significantly in every region assessed (Table 

4), with median slopes of -0.8 and -1.5 

μeq/L/year for the period 1990-2000. 

Nitrogen and NO3

-

 concentrations also 

declined significantly in every eastern 

region, with the exception of the Ridge and 

Blue Ridge provinces, where declines were 

not significant. These trends appear to be 

driven by a decline in NO3

-

 in deposition 

after 1995. Nitrogen and NO3

-

 wet 

deposition actually increased significantly in 

the Upper Midwest. 

Regional deposition trends for acidity and 

base cations. It is clear from the emission 

and deposition data that deposition closely 

follows emissions for sulfur, and perhaps for 

nitrogen. This is a positive result for 

evaluating the effectiveness of the CAAA. 

The expectation is that these declines in acid 

anions should result in a decline in acidity in 

precipitation. However, most studies have 

failed to find a commensurate decline in 

acidity (hydrogen ion) corresponding to the 

decline in SO4

2- deposition, although the 

declines in hydrogen ion are statistically 

significant at many NADP stations (Lynch 

et al. 2000). Some investigators have 

attributed the lower than anticipated 

response in hydrogen ion to decreased 

deposition of base cations that can offset 

declines in the wet deposition of sulfate 

(Hedin et al. 1994, Lynch et al. 2000).
26

Figure 10. Cumulative frequency distribution of slopes for SO4

2- (panel a, top) and nitrogen (panel b, 

bottom) trends in wet deposition, from regional NADP/NTN data for the period 1990-2000. The 

x-axis is the slope (rate of change in μeq/l/yr). The y-axis is the cumulative percent of the 

population; the median rate of change for each region corresponds to the value of the x-axis at the 

50th percentile of the curve. 
27

Table 4. Regional trend results (1990-2000) for atmospheric deposition (wet-only annual concentration 

data from NADP/NTN network) in acid sensitive regions. All units are μeq/L/yr. Values are the 

median slopes for each region, with significance determined by calculating confidence intervals 

around each regional median. 

Region SO4

2- Nitrogen NO3- Base Cations Hydrogen 

New England -0.96** -0.26* -0.20** +0.02ns -0.81** 

Adirondacks -1.47** -0.37** -0.38** +0.01ns -1.48** 

Appalachian Plateau -1.55** -0.41** -0.37** +0.01ns -1.54** 

Ridge/Blue Ridge Province -0.96** -0.16ns -0.13ns +0.05ns -0.94ns

Upper Midwest -0.75** +0.24** +0.21** +0.18** -0.57** 

ns regional trend not significant (p > 0.05) 

* p < 0.05 

** p < 0.01 

We evaluated the deposition patterns for 

base cations (Ca2+ plus Mg2+) and hydrogen 

ion in each region. In watersheds, base 

cations are weathering by-products of the 

neutralization reactions that produce ANC. 

It is not well understood if base cation 

deposition from the atmosphere contributes 

quantitatively important amounts of cations 

to soils, or to the neutralization capacity of 

the watershed. A long-term decline in the 

deposition of base cations (Hedin et al. 

1994) has been suggested as a possible 

cause of declining base cation 

concentrations in surface waters (Driscoll et 

al. 1989). Widespread declines in surface 

water base cations have offset decreases in 

SO4

2- concentrations, and minimized or 

prevented recovery in ANC in many waters 

(e.g., Stoddard et al. 1999). While median 

slopes for base cation deposition trends were 

positive in every eastern region (Figure 

11b), only in the Upper Midwest did wet 

deposition concentrations of base cations 

increase significantly. 

The combination of significantly decreasing 

SO4

2- deposition (and to a lesser extent 

declining nitrogen) and largely unchanged 

base cations deposition, as reported in this 

analysis, leads to an expectation of declining 

acidity in wet deposition as well. In fact, wet 

deposition of hydrogen ion has decreased in 

every region (Figure 11a), and is significant 

in every region except the Ridge and Blue 

Ridge provinces (Table 4). This is contrary 

to previous analyses (e.g., Lynch et al. 2000) 

using data through 1997, and can be 

attributed to the lack of trend in base cation 

concentrations for the period 1990-2000.
28

Figure 11. Distribution of slopes for acidity (hydrogen ion; panel a, top) and base cations (Ca2+ +Mg2+) 

(panel b, bottom) trends in regional NADP/NTN data, by region, for the period 1990-2000. 
29

What is ‘background’ precipitation 

chemistry? Miller (1999) analyzed the 

relationship between NADP/NTN SO4

2- wet 

deposition and sulfur emission inventories 

for the period 1980-98, using data from New 

England. We used his equations for the 

relationship at inland NADP/NTN stations 

to extrapolate SO4

2- deposition to a 

hypothetical condition of zero SO2 

emissions. The predicted range of 

precipitation SO4

2- concentrations at zero 

SO2 emission was 1.9 to 4.4 kg/ha at the six 

stations for which Miller developed 

equations. The mean SO4

2- deposition 

intercept (i.e. at zero emissions), was 2.9 

kg/ha. Assuming consistent climatic 

conditions, this deposition would equate to 6 

to 10 μeq/L SO4

2- in precipitation as 

‘background’, based on current 

concentration-deposition relationships. 

A background concentration of 6 to 10 

μeq/L can be compared to the present range 

of 20 to 60 μeq/L of wet SO4

2- at 

NADP/NTN sites in the eastern U.S., and 5 

to 20 μeq/L in the western U.S. The 

precipitation pH represented by this 

background SO4

2- would be pH 5.0 to 5.2, 

not including acidity from nitrate or organic 

acids. This analysis supports the inference of 

Charlson and Rodhe (1982) that the preindustrial background precipitation pH in 

the eastern U.S. was around 5.0, setting the 

baseline against which to compare modern 

chemistry. When other sources of acidity are 

included, it is unlikely that regulatory 

actions will be able to increase precipitation 

pH in the eastern U.S. to mean values 

greater than pH 5.0. 

Regional trends in precipitation amount. 

Climatic variation can be a major 

complication for interpreting chemical 

changes in surface waters. In particular, 

more dilute concentrations of both acid 

anions and base cations can be expected to 

occur in wet years, and lower concentrations 

may result during periods of drought, at least 

in streams and drainage lakes. In 

Midwestern seepage lakes, drought may 

have the effect of isolating lakes from baserich groundwater, essentially creating 

temporarily perched seepage lakes, and 

short-term acidification (Webster et al. 

1990). During the period investigated in this 

report, annual rainfall was relatively uniform 

in all of the regions (Figure 12). None of the 

trends in precipitation amount were 

significant for the 1990-2000 time period, 

and we have not considered trends in rainfall 

as part of our analysis of surface water 

trends. Seasonal wet and dry spells may 

none-the-less impart extra noise to the data, 

but we assume that the length of the data 

record smoothes the noise relative to 

seasonal variation. In short, we do not 

consider that climate variation is a major 

factor affecting the trends in the east and 

northeast with one exception: recovery from 

the drought of the 1980s in the Upper 

Midwest (Webster and Brezonik 1995) is 

believed to have influenced lake chemistry 

in the early 1990s, as discussed later in this 

document. 

RESPONSE OF SURFACE WATERS, 

1990-2000 

Important response variables. Our 

analysis of surface water response to 

changing deposition focuses on the key 

variables that play major roles in 

acidification and recovery: 

1) SO4

2-

 and NO3

-

, the acid anions that, 

when found in surface waters, are 

often the indicators of acidic 

deposition. Trends in the 

concentrations of these anions are 

inferred to reflect recent trends in 

deposition (especially SO4

2–) and in 

ecosystem response to long-term 

deposition (e.g., NO3

–

).
30

Figure 12. Annual precipitation in the study regions (National Atmospheric Deposition Program/National 

Trends Network 2002). Symbols represented mean values for sites in each region. Locations of 

NADP/NTN sites in each region listed in Table 2. 
31

Figure 13. The relationships between Gran ANC and [Ca2+ + Mg2+], and between Gran ANC and [K+ + 

Na+

] in acid-sensitive regions. 
32

2) Base cations, the by-products of the 

weathering reactions that neutralize 

acids in watersheds. We have 

chosen to evaluate the response of 

the two divalent cations, calcium and 

magnesium (henceforth [Ca2+ + 

Mg2+]) because of the nature of their 

behavior in soils; these two cations 

are quantitatively the most important 

for generation of Gran ANC (Figure 

13). In most regions, the relationship 

between [Ca2+ + Mg2+] and Gran 

ANC is much stronger than between 

potassium (K+

) plus sodium (Na+). 

The addition of [K+

 + Na+] to the 

cation-ANC relationship imparts 

extra variability to the interpretation 

of base cation response, although the 

patterns differ by region. 

3) pH and Gran ANC, the measures of 

acidity. These variables reflect the 

outcome of interactions between 

changing acid anions and trends in 

base cations. 

4) Aluminum, one of the most 

important variables for biological 

impacts due to its toxicity in the 

ionic form. Changes in 

concentrations and forms of 

aluminum are largely driven by 

changes in pH and DOC, with lower 

and less toxic aluminum 

concentrations resulting from high 

pH values and DOC concentrations. 

5) Concentrations of Dissolved 

Organic Carbon (DOC), as a 

surrogate for organic acidity. 

Organic acids are additional, largely 

natural, sources of acidity in surface 

waters. 

Representative data from the regions. The 

trends we report below are for the period of 

focus for this report, 1990-2000. Many of 

the monitoring sites for which we report 

data have been operating since the early 

1980s, and examining the entire period of 

record helps to put current trends into a 

longer-term context. As detailed earlier, 

SO4

2– deposition has been declining in the 

northern and eastern U.S. since the 1970s, 

and changes produced by the CAAA of 

1990 overlay the changes already underway. 

In order to illustrate this historical context, 

we have chosen representative sites from 

each region, and present time series of the 

important response variables outlined above. 

‘Representative’ in this case refers to the 

observed time trends; each site exhibits 

trends in the 1990s that are typical of the 

region. 

Salmon Pond in eastern Maine (Figure 14) is 

one of 24 drainage lakes sampled as part of 

the LTM project in New England which 

operates sites in Maine and Vermont. 

Salmon Pond has been sampled since early 

1982, pre-dating the beginning of the LTM 

project by a year. It is sampled quarterly, in 

the spring, summer and fall. 

Sampling at Darts Lake (Figure 15) in the 

Adirondacks began with the RILWAS 

project (Gherini et al. 1989) in 1982; it 

became part of the LTM project in 1983, 

and has been sampled monthly through the 

1980s and 1990s. 

The East Branch of the Neversink River 

(Figure 16) is a famous trout fishing stream 

in the Catskill Mountains of New York, and 

exhibits trends typical of the Northern 

Appalachian Plateau in the 1990s. It has 

been sampled as part of the LTM project 

since 1984, with variable sampling 

frequencies. The Neversink is sampled for 

both baseflow and episodic chemistry. 

Vandercook Lake in Wisconsin (Figure 17) 

exhibits trends representative of the Upper 
33

Midwest. This seepage lake has been 

sampled quarterly in the spring, summer, 

and fall since 1984. 

The Staunton River, showing trends 

representative of the Ridge/Blue Ridge 

region of Virginia (Figure 18), is located in 

 Shenandoah National Park. It became part 

of the LTM network in 1992, but has been 

sampled as part of various programs funded 

by the National Park Service since 1987. It 

is currently sampled on a schedule similar to 

the Neversink River. 
34

Figure 14. Time series data for SO4

2–, NO3–

, base cations [Ca2+ + Mg2+], Gran ANC, pH, and DOC in 

Salmon Pond, Maine (New England region). Significant trends are indicated by trend lines. 

Shaded box indicates time period of analyses reported here. 
35

Figure 15. Time series data for SO4

2–, NO3–

, base cations [Ca2+ + Mg2+], Gran ANC, pH, and DOC in 

Darts Lake, NY (Adirondack region). Significant trends are indicated by trend lines. Shaded 

box indicates time period of analyses reported here. 
36

Figure 16. Time series data for SO4

2–, NO3–

, base cations [Ca2+ + Mg2+], Gran ANC, pH, and DOC in the 

E. Branch of the Neversink River, New York Catskills (Northern Appalachian Plateau region). 

Significant trends are indicated by trend lines. Shaded box indicates time period of analyses. 
37

Figure 17. Time series data for SO4

2–, NO3–

, base cations [Ca2+ + Mg2+], Gran ANC, pH, and 

DOC in Vandercook Lake, Wisconsin (Upper Midwest region). Significant trends are indicated 

by trend lines. Shaded box indicates time period of analyses reported here. 
38

Figure 18. Time series data for SO4

2–, NO3–

, base cations [Ca2+ + Mg2+], Gran ANC, pH, and 

DOC in the Staunton River, Virginia (Ridge/Blue Ridge region). Significant trends are indicated 

by trend lines. Shaded box indicates time period of analyses reported here. 
39

Trends in sulfate by region. SO4

2– declined 

in surface waters in the glaciated regions of 

the North and East by median values 

between –2 and –4 μeq/L/year (Figure 19), 

with the smallest changes in New England 

and the largest declines in the Upper 

Midwest. The only exception to the pattern 

of declining SO4

2– was in the Ridge/Blue 

Ridge provinces, which by virtue of their 

soil characteristics were not expected to 

show a rapid response to changes in 

deposition (Cosby et al. 1986, Church et al. 

1989). The median SO4

2– increased slightly 

in Ridge and Blue Ridge streams. All of the 

regional SO4

2– trends are highly significant 

(Table 5), and are consistent with the trends 

reported previously, including other regions 

of North America and Europe (Mattson et al. 

1997, Stoddard et al. 1999, Evans and 

Monteith 2001, Skjelkvåle et al. 2001). 

The marked declines in SO4

2– concentrations 

in the glaciated portions of the North and 

East are almost certainly direct responses to 

declining emissions and SO4

2– deposition in 

the 1990s, and represent the most dramatic 

effects of Title IV of the CAAA and 

previous emissions regulations. These 

changes in emissions and deposition 

continue the trend in declining SO4

2– that 

has been occurring for almost 3 decades 

(Stoddard et al. 1999). 

The small increase in SO4

2– concentrations 

in the Ridge and Blue Ridge provinces 

results from the SO4

2– adsorption properties 

typical of soils found throughout this 

unglaciated region, which have the effect of 

both decoupling (in time) trends in 

deposition and surface waters (Tables 4 and 

5), and lowering the ambient concentrations 

of SO4

2– well below those found in 

deposition (Figure 7). The future response of 

SO4

2– concentrations in this region will be a 

critical element in governing any potential 

recovery in this region.

Table 5. Regional trend results for Long-Term Monitoring sites for the period 1990 through 

2000. Values are median slopes for set of sites in each region. Units for sulfate, nitrate, base 

cations [Ca2+ + Mg2+], Gran ANC and hydrogen are μeq/L/year. Units for DOC are mg/L/year. 

Units for aluminum are μg/L/year. 

Region SO4

2– NO3– Base 

Cations

Gran 

ANC Hydrogen DOC Aluminum

New England Lakes -1.77** +0.01ns -1.48** +0.11 ns -0.01ns +0.03* +0.09ns

Adirondack Lakes -2.26** -0.47** -2.29** +1.03** -0.19** +0.06** -1.12** 

Appalachian Streams -2.27* -1.37** -3.40** +0.79* -0.08* +0.03ns +0.56ns

Upper Midwest Lakes -3.36** +0.02ns -1.42** +1.07** -0.01* +0.06** -0.06ns

Ridge/Blue Ridge Streams +0.29** -0.07** -0.01ns -0.07ns +0.01ns NA NA 

ns regional trend not significant (p > 0.05) 

* p < 0.05 

** p < 0.01 

NA insufficient data 
40

Figure 19. Cumulative frequency diagram (distribution) of slopes (μeq/L/year) for SO4

2–

concentrations in surface water monitoring sites, by region for the period 1990-2000.
41

Figure 20. Cumulative frequency diagram (distribution) of slopes (μeq/L/year) for NO3

–

concentrations in surface water monitoring sites, by region for the period 1990-2000. 

Trends in nitrate by region. In general, 

trends in NO3

–

 were much smaller than 

trends in SO4

2–, with the only ecologically 

significant changes occurring in the two 

regions with the highest ambient NO3

–

concentrations (Figure 20). Both lakes in 

the Adirondacks and streams in the Northern 

Appalachian Plateau exhibited small but 

significant downward trends in NO3

–

 in the 

1990s (Table 5). Both of these regions are 

central to the debate over whether nitrogen 

saturation is a legitimate threat to the health 

of forests and surface waters (Stoddard 

1994, Aber et al. 1998). While declining 

NO3

–

 concentrations in these regions are a 

positive development for these ecosystems, 

we clearly do not know if these trends will 

continue, especially since they do not reflect 

recent trends in emissions or deposition 

(Table 4, Figures 7-9). The presence of 

strong upward trends in NO3

–

 in these same 

regions in the 1980s (Murdoch and Stoddard 

1992, Stoddard 1994) suggests that trends 

on the scale of a single decade probably 

represent noise (variability) in a long-term 

pattern of changing NO3

–

 leakage from 
42

forested watersheds controlled by factors 

that will take many years to determine. 

While great uncertainty exists, and the time 

scales of nitrogen saturation may be longer 

than previously considered (e.g., centuries 

rather than decades), the long-term retention 

nitrogen deposition in forested regions is 

unlikely to continue indefinitely (Aber et al. 

in press). 

In New England and the Upper Midwest, 

where ambient NO3

–

 concentrations are 

much lower than in the Adirondacks and 

Northern Appalachian Plateau (Figures 14-

17), NO3

–

 concentrations in surface waters 

were unchanged. The Ridge/Blue Ridge 

province registered a small, but significant, 

decrease in NO3

–

 during the 1990s, but 

interpretation of trends for NO3

–

 in this 

region is complicated by an outbreak of 

Gypsy Moths that also occurred during this 

period. Forest defoliation by Gypsy Moths 

was the cause of a pulse in NO3

–

 export 

from many streams in this region in the mid1990s (Eshleman et al. 1998). 

Trends in Gran ANC and acidity, by 

region. In a very real sense, increasing 

trends in Gran ANC would be the strongest 

signs of recovery from acidification in 

response to the CAAA. There was a 

tendency toward increasing Gran ANC 

(Figure 21) in all of the glaciated regions of 

the North and East (i.e., New England, 

Adirondacks, Northern Appalachian Plateau 

and Upper Midwest), and decreasing Gran 

ANC in the Ridge/Blue Ridge province. 

Only the regional increases in the 

Adirondacks, Northern Appalachian Plateau, 

and Upper Midwest were significant (Table 

5). Median increases of ca. +1 μeq/L/year 

in the Northern Appalachian Plateau, 

Adirondacks and Upper Midwest represent 

significant ecological recovery from 

acidification. 

Previous regional analyses, using data 

through 1995, have highlighted the lack of 

ANC recovery in the Adirondacks, Northern 

Appalachian Plateau and Upper Midwest 

(Stoddard et al. 1999). In the Adirondacks 

and Northern Appalachians, many of the 

sites exhibit the pattern shown for Darts 

Lake in Figure 15 and the Neversink River 

in Figure 16, with Gran ANC increasing 

after an inflection point in the early- to mid1990s. The current Gran ANC trends (Table 

5) in the Adirondacks and Northern 

Appalachian Plateau appear to have reversed 

the pattern of lack of recovery, with further 

acidification occurring in some cases, that 

occurred throughout the 1980s and early 

1990s in this region. 

In the Upper Midwest, the downward Gran 

ANC trends reported by Stoddard et al. 

(1999) for the period 1990-95 appear to 

have been a temporary departure from the 

long-term recovery of ANC in this region. 

The trends reported for the 1990s (Table 5) 

are consistent with the long-term pattern of 

increasing Gran ANC observed at individual 

sites. 

Stoddard et al. (1999) reported recovery in 

Gran ANC in the New England region using 

data through 1995. This trend is no longer 

evident when all data from the 1990s are 

examined (Table 5). The temporal pattern 

for New England lakes appears to be the 

reverse of that exhibited by lakes in the 

Adirondacks, with long-term recovery in 

Gran ANC having fallen off after a similar 

inflection point in the early 1990s (Figure 

14). The slight tendency for Gran ANC to 

increase in this region is not significant, with 

nearly as many lakes showing small 

decreases (11) as show slight increases (13). 

Gran ANC showed an insignificant 

downward trend in the Ridge/Blue Ridge 

province, consistent with the small increases 

in SO4

2– observed. It is clear that, while the 

evidence for a widespread recovery in ANC 

in the glaciated North and East is quite 
43

strong, no such evidence exists for the 

unglaciated Ridge/Blue Ridge province. 

Widespread changes in Gran ANC 

translated into some small, but significant, 

changes in acidity (hydrogen ion; Figure 

22). The largest significant decreases in 

hydrogen ion occurred in the Adirondacks 

(Table 5), but on an equivalent basis they 

were less than 20% of the changes observed 

in Gran ANC in this region. The moderate 

decrease observed in the Northern 

Appalachians was likewise small relative to 

the change in Gran ANC (< 10% of the 

ANC change). All of the other regions 

exhibited median slopes near zero, and only 

the Upper Midwest trend was significant 

(Table 5). To put these changes in hydrogen 

ion in context, a decline in hydrogen 

concentration of 1 μeq/L/year in an acidic 

lake with pH 4.7 would equate to a of pH 

5.00 over 10 years. All of the changes 

observed here are considerably smaller than 

this. 

Trends in base cations, by region. All of 

the glaciated regions on the northern and 

eastern U.S. exhibited declines in base 

cation [Ca2+ + Mg2+] concentrations in the 

range of –1.5 to –3.4 μeq/L/year (Figure 

23), and all of the regional trends were 

highly significant (Table 5). One of the most 

universal watershed responses to acidic 

deposition is the mobilization of base 

cations from soils (Galloway et al. 1983); as 

rates of acidic deposition decline in the 

North and East, and the supply of acid 

anions to watersheds soils decreases, the 

rates of cation mobilization are also 

expected to decrease. Lowered rates of 

cation mobilization translate to downward 

trends in surface water base cation 

concentrations, a change widely observed in 

the northern hemisphere for more than a 

decade. Keller et al. (2001) suggested that a 

trend of declining Ca2+ has been occurring 

for more than a century in Ontario lakes. 

As a broad generalization, SO4

2– has 

decreased at a rate of approximately –2.5 

μeq/L/year (the mean of regional median 

slopes), and NO3

– at a rate of –0.5 

μeq/L/year, in surface waters of glaciated 

terrain. These rates of change set an upper 

limit to our expectation of ANC recovery of 

+3 μeq/L/yr (i.e., the sum of SO4

2– and NO3–

trend magnitudes). The Gran ANC increase 

is actually about one-third of this maximum 

, +1 μeq/L/year. The difference between the 

observed Gran ANC trend and the maximum 

trend estimated from rates of acid anion 

change can be almost entirely explained by 

regional declines in base cations; the 

average regional median decline in [Ca2+ + 

Mg2+] was about –2.0 μeq/L/year.
44

Figure 21. Cumulative frequency diagram (distribution) of slopes (μeq/L/year) for Gran ANC 

concentrations in surface water monitoring sites, by region for the period 1990-2000. 
45

Figure 22. Cumulative frequency diagram (distribution) of slopes (μeq/L/year) for hydrogen ion 

concentrations in surface water monitoring sites, by region for the period 1990-2000. 
46

The decline in [Ca2+ + Mg2+], while not 

unexpected, limits the magnitude of ANC 

recovery by offsetting some of the decline in 

SO4

2– and NO3–.. Moreover, if the rate of 

base cation decline is equal to the rates of 

decline in acid anions, recovery will be 

prevented. This is the pattern observed in 

several regions of North America in 

previous regional analyses (Stoddard et al. 

1998a, Stoddard et al. 1998b) using data 

through the early 1990s. The observation 

that acid anions, and particularly SO4

2–, 

concentrations are decreasing at a faster rate 

than base cations in most regions (Table 5) 

leads to an increase in Gran ANC, and is 

consistent with the pattern that has been 

observed in northern Europe since about 

1990 (Stoddard et al. 1999). 

Several researchers (e.g., Likens et al. 1996, 

Lawrence et al. 1999, Driscoll et al. 2001) 

have proposed that the recovery of surface 

waters will be substantially delayed because 

soil impoverishment, induced by decades of 

base cation leaching by the anions in acidic 

deposition, has reached critical levels. At 

least one aspect of the current analysis 

argues against this hypothesis. The regional 

response of base cations to changing SO4

2–

deposition is following expectations (e.g., 

[Ca2+ + Mg2+] is decreasing at lower rates 

than SO4

2– in glaciated terrain, and is 

unchanged in unglaciated terrain) in every 

region. 

Trends in dissolved organic carbon 

(DOC), by region. Another potential factor 

influencing the responsive in ANC and 

acidity is the general increase in DOC 

(Figure 24). All regions with sufficient 

DOC data exhibited increases in 

concentrations, and all regional trends were 

significant with the exception of the 

Northern Appalachian Plateau (the region 

with the lowest median DOC 

concentrations). While the character of the 

DOC found in these watersheds has not been 

evaluated, some portion of the DOC is 

organic acid. The median increase of 0.05 

mg/L/year corresponds to a increase of 

about 10% region-wide, similar to trends 

reported elsewhere in the northern 

hemisphere (Evans and Monteith 2001, 

Skjelkvåle et al. 2001). 

It is beyond the scope of an analysis of 

monitoring data, like the one presented here, 

to determine the mechanism for DOC 

increase unequivocally. However, there are 

several aspects of the DOC trends that 

warrant further attention. First, increases in 

DOC have now been reported from nearly 

all regions of North America (this report) 

and Europe (Skjelkvåle et al. 2001). This 

argues for a large-scale cause, and both 

climate change (warmer temperatures are 

expected to increase decomposition, and 

could potentially increase the export of DOC 

from watersheds) and decreasing acidic 

deposition are possible hypotheses. Several 

researchers have proposed that at least part 

of the increase in mineral acids (e.g., SO4

2–

and NO3

–

) observed in acidifying 

watersheds in the past (Rosenqvist 1978, 

Krug and Frink 1983) resulted from the 

replacement of natural organic acidity with 

mineral acidity. They thus propose that the 

process of acidification was accompanied by 

a decrease in surface water DOC 

concentrations. This hypothesis has received 

some support from paleolimnological 

studies, where reconstructions of historic 

DOC concentrations show decreases during 

the period of acidification in the 1900s 

(Davis et al. 1985, Kingston and Birks 1990, 

Dixit et al. 2001). David et al. (1999) were 

also able to measure a decrease in organic 

anions in response to the experimental 

whole-watershed acidification experiment at 

the Bear Brook Watershed in Maine (Norton 

et al. 1999). In one of the few studies of 

DOC behavior during recovery from 
47

acidification, Wright et al. (1993) concluded 

that ANC increases in a small watershed 

where rates of acidic deposition were 

experimentally reduced were limited by the 

increasing role of organic acids, and an 

overall increase in DOC.

Figure 23. Cumulative frequency diagram (distribution) of slopes (μeq/L/year) for base cations 

[Ca2+ + Mg2+] in surface water monitoring sites, by region for the period 1990-2000. 
48

Figure 24. Cumulative frequency diagram (distribution) of slopes (mg/L/year) for DOC 

concentrations in surface water monitoring sites, by region for the period 1990-2000. The 

Ridge/Blue Ridge province does not have sufficient DOC data to allow trend analysis. 
49

In a regional study of trends, like the one 

presented here, one might predict that, if this 

hypothesis (the replacement of organic 

acidity by mineral acidity during 

acidification) is correct, then the current 

rates of DOC increase should be correlated 

to rates of SO4

2– decline. This is, in fact, 

observed in the trend data for all lake 

regions (Figure 25). The largest rates of 

DOC increase are observed at the sites with 

the strongest declines in surface water 

SO4

2–. This relationship is strongest in high 

DOC systems (i.e., those with DOC > 5 

mg/L), but is also observed in low DOC 

lakes.

Figure 25. Relationship between trends in DOC and trends in SO4

2– concentrations for lakes. This graph 

includes data from the Maine high elevation and seepage lakes in order to produce a relationship 

with sufficient numbers of high DOC in small, responsive lakes. 
50

Figure 26. Cumulative frequency diagram (distribution) of slopes (μg/L/year) for aluminum 

concentrations in surface water monitoring sites, by region for the period 1990-2000. Data from the 

Adirondacks are for inorganic monomeric aluminum; all others are total dissolved aluminum. 
51

Trends in aluminum by region. Aluminum 

plays a major role in the biotic 

impoverishment that often accompanies 

surface water acidification (e.g., Baker and 

Schofield 1982, Cleveland et al. 1986). 

Decreasing trends in aluminum, particularly 

in the most toxic form of aluminum 

(inorganic monomeric aluminum), would 

therefore be a biologically-relevant indicator 

of recovery. Only one of the regions 

analyzed here exhibited regional decreases 

in aluminum (Figure 26). In the 

Adirondacks, the median decline in 

inorganic monomeric aluminum was 

approximately –1.0 μg/L/year, a very small 

number in terms of importance to biota. 

However, this change is consistent with 

regional increases in Gran ANC and 

decreases in hydrogen ion. The Adirondack 

trends are also consistent with reported 

decreases in toxic aluminum in soil water 

samples from the Hubbard Brook 

Experimental Forest in New Hampshire 

(Palmer and Driscoll 2002). No regional 

trends other than the Adirondacks were 

significant (Table 5). 

Our analysis for aluminum is complicated 

by the fact that the most consistently 

collected data for aluminum are for total 

dissolved aluminum, which includes both 

organic forms and polymeric inorganic 

forms. Small changes in inorganic 

monomeric aluminum, which may make up 

only a small portion of total dissolved 

aluminum, may be undetectable if the other 

forms are relatively unchanged, or are 

highly variable. We do not therefore 

consider the general lack of trends in 

regional aluminum concentrations 

conclusive, although they are generally 

consistent with patterns of unchanged 

hydrogen ion. 

Relationship between ANC trends and 

ANC class: It is reasonable to expect, based 

on our conceptual understanding of the 

process of acidification, that decreasing rates 

of acidic deposition, and decreasing surface 

water concentrations of acid anions, will 

produce the greatest rates of recovery in the 

sites that have undergone the most severe 

acidification. One outcome of this 

expectation would be larger upward trends 

in acidic lakes and streams, small upward 

trends in low ANC sites, and no change in 

the relatively insensitive sites we classify as 

moderate ANC. An analysis of Gran ANC 

trends by ANC class, designed to test this 

expectation is shown in Table 6. Using data 

from all of the sites in regions where 

decreases in surface water SO4

2– and/or 

NO3

–

 have occurred, we find that acidic 

lakes and streams exhibit a highly 

significant median increase in Gran ANC of 

+1.3 μeq/L/year. Low ANC sites show a 

smaller, but none-the-less significant, 

median ANC increase of +0.8 μeq/L/year. 

Moderate ANC sites, those with mean ANC 

values greater than 25 μeq/L, show no 

significant change in Gran ANC.
52

Table 6. Slopes of trends in Gran ANC in acidic, low ANC and moderate ANC sites for the period 1990-

2000. Analysis includes all sites in New England, Adirondacks, Appalachian Plateau and Upper 

Midwest; Ridge and Blue Ridge sites excluded. 

ANC Class Number of 

Sites 

Change in 

Gran ANC (μeq/L/yr) 

Acidic 

(ANC < 0 μeq/L) 26 +1.29** 

Low ANC 

(0 < ANC < 25 μeq/L) 51 +0.84** 

Moderate ANC 

(25 < ANC < 200 μeq/L) 43 +0.32 ns

ns trend not significant (p > 0.05) 

* p < 0.05 

** p < 0.01 

Regional trends in probability networks. 

The data described thus far in this report 

come from the LTM project, and from other 

monitoring networks with similar objectives. 

The method of site selection utilized for 

LTM varied among regions, and was in part 

driven by the existence of historical data 

and/or ongoing data collection, rather than 

statistical frameworks to determine 

population trends. Stoddard et al. (1998b) 

placed LTM sites from the Northeast in the 

lake classification system of Young and 

Stoddard (1996) and found that they largely 

covered the range in current status (e.g., 

current Gran ANC and base cation 

concentrations) of the most sensitive 

subpopulation(s) of lakes in the region. The 

trend results, however, indicated two 

different classes of response: lakes in New 

England were increasing in ANC (based on 

data for the period 1984-94), while those in 

the Adirondacks were acidifying slightly, 

even when the current status data indicated 

they belonged in the same lake 

subpopulation. 

The data in this report offers the first 

opportunity to extend the analysis of LTM 

representativeness into the realm of trends, 

as well as changing population status. By 

comparing the magnitudes of trends 

measured from repeated probability surveys 

(i.e., from the TIME project) with those 

observed in the same regions through LTM 

sampling, we have an opportunity to verify 

the validity of LTM trends. 

In both the Adirondacks and New England, 

population-wide results inferred from TIME 

data show very similar trends in SO4

2– to 

those for LTM sites (Figure 27). Both 

datasets suggest highly significant regional 

decreases in SO4

2– slightly larger than -2 

μeq/L/year in the Adirondacks and -1.8 

μeq/L/year in New England (Tables 5 and 

7). 

Both datasets suggest similar ranges of trend 

slopes for Gran ANC (Figure 28). In the 

Adirondacks, regional increases in Gran 

ANC are significant (Tables 5 and 7) in both 

the LTM sites and TIME population data. 

In New England, only the population trend 

in Gran ANC was significant (Table 7), 

although both datasets suggest that New 

England lakes are undergoing slight 
53

increases in Gran ANC. The greater 

response in TIME lakes may be due to the 

different length of record between the 

networks. If increases in Gran ANC 

accelerated in the late 1990s, as they did in 

the Adirondacks, then the TIME data may 

well estimate a larger rate of change than 

LTM data.

Table 7. Regional trend results for populations of sites in acid sensitive regions. Results from 

TIME probability sites are extrapolated to regional target populations. Lake trends in the 

Adirondacks and New England are for the period 1991-2001. Stream trends in the Northern 

Appalachian Plateau are for the period 1993-2001. Values are weighted median slopes for all 

lakes or streams in each region. Units for SO4

2–, NO3–

, base cations [Ca2+ + Mg2+], Gran ANC 

and hydrogen are μeq/L/yr. Units for DOC are mg/L/yr. Units for aluminum (inorganic 

monomeric) are μg/L/yr. 

Region SO4

2– NO3– Base 

Cations

Gran 

ANC Hydrogen DOC Aluminum

Adirondack Lakes -2.10** +0.01ns -1.22* +0.56* -0.09ns +0.09* +0.66ns

New England Lakes -1.88** +0.02ns -1.57** +0.40* +0.01ns +0.08* -1.94ns

Appalachian Streams -0.64* +0.04ns -0.32ns +0.34* -0.01ns +0.01ns +0.14ns

ns regional trend not significant (p > 0.05) 

* p < 0.05 

** p < 0.01 
54

Figure 27. Comparison of trends in SO4

2– in TIME probability sites (expressed as the distribution 

of slopes in the target population) and LTM sites in New England and the Adirondacks.
55

Figure 28. Comparison of trends in ANC in TIME probability sites (expressed as the distribution 

of slopes in the target population) and LTM sites in New England and the Adirondacks. 
56

TIME population estimates and LTM data 

also lead to the same conclusions for 

regional trends in [Ca2+ + Mg2+]

concentrations in the two regions (Figure 

29). In New England, the distributions of 

slopes are nearly identical, with median 

decreases of -1.6 μeq/L/year in the TIME 

population estimate and -1.5 μeq/L/year the 

LTM data; both regional trends are highly 

significant (Tables 5 and 7). In the 

Adirondacks, both estimates are significant, 

but the estimate based on TIME population 

data (-1.2 μeq/L/year) is only half that of the 

LTM data (-2.3 μeq/L/year). The reason for 

this modest discrepancy again may be the 

different time periods represented by the 

data.

Overall, the conclusions drawn from either 

TIME probability sampling and LTM 

intensive sampling are the same. In New 

England lakes, SO4

2– is declining only 

slightly more than base cations, and the 

result is only a small increase in Gran ANC 

(significant in the TIME data; insignificant 

for LTM). In the Adirondacks, base cation 

declines exceed SO4

2– declines by a large 

amount, resulting in more substantial 

recovery in Gran ANC. 

Do trends in deposition translate into 

trends in surface waters? A major goal of 

this assessment report is to evaluate the 

changes in surface water chemistry in 

relation to CAAA reductions in deposition. 

In particular, did the changes in deposition 

result in commensurate changes in surface 

water chemistry in the past decade? There 

has been discussion in the scientific 

literature about the recovery response for 

SO4

2– in surface waters: would the response 

be rapid or delayed (Church et al. 1989, 

Church 1999)? 

It is not possible to compare directly 

changes in SO4

2– concentrations in surface 

waters and deposition, because of the poorly 

quantified effects of dry deposition and 

evapotranspiration. Both dry deposition and 

the evaporative concentration of ions in 

surface waters cause SO4

2– concentrations in 

lakes and streams to be typically 2-3 times 

higher than in deposition (Sisterson et al. 

1990, Rustad et al. 1994). Thus, higher 

concentrations in surface waters lead to the 

likelihood of greater downward slopes for 

SO4

2– concentrations in surface waters than 

in deposition. In contrast, the percent 

change in SO4

2– should be relatively 

unaffected by ambient concentrations, 

assuming that dry deposition declines at the 

same rate as wet deposition. In this section, 

we present a comparison of the slopes of 

change in both deposition and surface 

waters, using the regional NADP/NTN and 

LTM data already presented. 

In New England, the Adirondacks and the 

Northern Appalachian Plateau, percentage 

decline in deposition of sulfate (expressed as 

%SO4

2–) are generally steeper than in 

surface waters, suggesting that most aquatic 

systems are exhibiting a lagged response. 

Apparently watershed soils retain enough 

atmospheric SO4

2– to reduce the rate 

declines occurring in deposition (Figure 30). 

Interestingly, the lakes and streams with the 

steepest declines in %SO4

2– in these regions 

have very similar rates to those in 

deposition, suggesting that the most 

responsive watersheds are acting essentially 

as flow-through systems. As expected, there 

is little correspondence between rates of 

%SO4

2– in streams and deposition in the 

Ridge and Blue Ridge provinces (Figure 

30), due to the adsorptive capacity of the 

highly weathered soils in the region.
57

Figure 29. Comparison of trends in base cations [Ca2+ + Mg2+] in TIME probability sites 

(expressed as the distribution of slopes in the target population) and LTM sites in New England 

and the Adirondacks. 
58

Figure 30. Comparison of % change in SO4

2– concentration in wet deposition and surface waters, 

by region. 
59

In the upper Midwest, the rate of decline in 

lakes is greater than the decline in 

deposition, probably reflecting the residual 

effects of the drought of the late 1980s. All 

of the lakes with strong negative trends in 

%SO4

2– are seepage lakes with long water 

residence times. These lakes had trends in 

sulfate that reflected a decline from the 

evaporative concentration that would have 

occurred during the drought, and the 

flushing of sulfur likely stored in wetlands 

during the drought (Dillon et al. 1997, 

Jeffries et al. 2002). Some SO4

2–may also be 

stored in the sediments of these lakes via 

sulfate reduction (Baker et al. 1986), 

although the rate of sulfate reduction is 

presumed to decrease as the concentrations 

decrease. 

We conclude that surface waters have 

responded relatively rapidly to the decline in 

sulfate deposition. There is reason to expect 

that additional reductions in deposition 

would result in additional declines in surface 

water concentrations of sulfate. 

Has there been a decline in the number of 

acidic waters? Such probability samples as 

TIME lakes allow us to estimate the number 

or proportion of surface waters in each 

region that are acidic. Because we have 

time series data in TIME, we have a unique 

opportunity to assess the changes in the 

number of acidic systems during the past 

decade. We have combined analyses of the 

TIME and LTM datasets to conduct this 

analysis, and have used data from the 

NSWS, conducted in the mid 1980s, to 

estimate the extent of acidification in the 

1980s in regions where TIME does not 

operate. 

New England Lakes: During 1991-94, 

TIME population data indicated that there 

were 386 lakes in New England with Gran 

ANC<0 μeq/L, or 5.6% of the population. 

Both TIME and LTM data from this region 

suggest that only a small increase in Gran 

ANC has occurred since that time period; 

combined TIME/LTM results suggest a 

change of ca. +0.3 μeq/L/year (Tables 5 and 

7). If we extend this time trend for 10 years, 

then sensitive lakes can be expected to have 

increased in ANC by ca. 3 μeq/L. In the 

1991-94 TIME probability survey there 

were 12 lakes with Gran ANC values 

between 0 and –3 μeq/L, and only these 

lakes are estimated to have become nonacidic by the present time (this estimate is 

for the time period 2001-04, 10 years after 

the original TIME estimate). We find little 

evidence of a profound change in the acidity 

status of this region, with the proportion of 

acidic lakes decreasing only from 5.6% to 

5.5% over the previous 10 years (Table 8). 

Adirondack Lakes: TIME population 

estimates from 1991-94 indicate that 13.0% 

of lakes, or 238 lakes, were acidic in the 

early 1990s. If we apply an approximate 

rate of change in Gran ANC of +0.8 

μeq/L/year to these estimates (based on 

trend slopes for TIME and LTM data in this 

region, Tables 5 and 7), then sensitive lakes 

could be expected to have increased their 

ANC values by 8 μeq/L during the 1990s. 

We estimate that at the present time there 

are ca. 149 Adirondack lakes with Gran 

ANC<0, or 8.1% of the population. This 

represents a decrease of nearly 5%; roughly 

38% of lakes that were acidic in the early 

1990s are estimated to be non-acidic a 

decade later. 

Northern Appalachian Plateau streams:

Population estimates from TIME surveys of 

streams in the Northern and Central 

Appalachian Plateau ecoregions 

(synonymous with the northern Appalachian 

Plateau region covered by the LTM project) 

indicate that 5,014 kilometers of streams 

were acidic in 1993-94 (at the beginning of 

TIME sampling); this represents 11.8% of 

the stream length. The approximate rate of 
60

change in Gran ANC in the region (Tables 5 

and 7) is +0.7 μeq/L/year, and extrapolated 

for a 10 year period this amounts to an 

increase in sensitive streams of +7 μeq/L. 

We estimate that roughly 3,600 kilometers 

of stream (or 8.5%) remain acidic in this 

region at the present time. This represents a 

ca. 28% decrease in acidic stream length 

over the decade. 

Upper Midwest Lakes: The TIME project 

has not operated in the Upper Midwest, so 

the best population estimates for this region 

come from the 1984 Eastern Lake Survey 

(Linthurst et al. 1986b). ELS sampling in 

the Upper Midwest estimated that 251 lakes, 

or 2.9% of the population, were acidic in 

1984 (Baker et al. 1991). LTM data suggest 

a rate of change of +1 μeq/L/year in this 

region (Table 5); extrapolated to the present, 

this represents an increase of +18 μeq/L of 

Gran ANC in sensitive lakes between 1984 

and 2002. In 1984, 80 lakes had Gran ANC 

values less than -18 μeq/L, and all of these 

are now estimated to be non-acidic. This 

represents a change from 2.9% acidic in 

1984, to 0.9% in 2002, an overall reduction 

in the number of acidic lakes of 68%. 

Ridge/Blue Ridge Province streams: While 

we do not have adequate data from TIME to 

calculate trends in this region, the data are 

adequate to estimate the number of acidic 

streams. TIME data indicate that fewer than 

400 km of streams (< 1%) were acidic in the 

Ridge and Blue Ridge region in 1993-94. 

Because we have no evidence of a 

significant change in Gran ANC in this 

region (Table 5), we estimate that <1% of 

Ridge/Blue Ridge streams remain acidic in 

2002. There has probably been no change in 

the number of acidic waters in this region in 

the past decade. 

Two important caveats in this analysis 

should be noted. First, as shown in Table 6, 

the rates of ANC increase (outside of the 

Ridge and Blue Ridge provinces) are more 

rapid in acidic lakes and streams than in low 

ANC sites. Due to sample size constraints (it 

is not possible to do the ANC class analysis 

by region) the changes in Gran ANC used in 

the analysis in Table 8 are based on the 

median change in all sites in a region. If 

acidic sites are recovering more rapidly than 

the population of sites as a whole, then the 

estimates of change in the number of acidic 

lakes and streams presented here will be 

conservative. 

It is also important to realize that the trends 

we report here are for recovery from chronic 

acidification. Our regression analysis 

focuses on the central tendency of sites (e.g., 

their median or mean ANC over time), and 

whether this central tendency is increasing 

or decreasing. We know that most sites 

exhibit seasonally lower ANC and pH 

values than would be captured by the trend 

analysis. In many cases, sites that are not 

chronically acidic none-the-less undergo 

short-term episodic acidification during 

spring snowmelt, or during intense rain 

events. In order to estimate the relative 

importance of episodic acidification in the 

regions we cover in this report, we examined 

the mean summer Gran ANC at each site, 

and compared it to the average minimum 

ANC experienced each spring during the 

period 1990-2000. The results of this 

analysis are shown in Figure 31. On 

average, spring ANC values in New 

England, the Adirondacks and Northern 

Appalachian Plateau are 30 μeq/L lower 

than summer values. This implies that lakes 

and streams in these regions would need to 

recover to Gran ANC values above 30 μeq/L 

before they could be expected not to 

experience acidic episodes. 

Estimating the magnitude of historical 

lake acidification. It is widely established 

that many present-day acidic lakes were at 

least marginally acidic in pre-industrial 

times (Davis and Anderson 1985, Sullivan 

1990, Cumming et al. 1994, Smol et al. 
61

1998). The main method for determining 

pre-historical pH is using siliceous fossil 

remains of diatom and chrysophyte algal 

species in lake sediment cores. Cumming et 

al. (1994) describes a classification system 

for the low ANC lakes with four types of 

response: 1) lakes that have not acidified; 2) 

lakes that acidified in the mid-1900s; 3) low 

pH lakes that acidified around 1900; 4) 

naturally acidic lakes that only acidified 

slightly. Two examples of these 

acidification responses are shown (Figure 

32) from New England. Little Long Pond, 

with a present Gran ANC of 10 μeq/L, 

acidified by about 0.2 pH units after 1920. 

Its neighbor Mud Pond (Gran ANC today of 

–20 μeq/L) acidified from about pH 5.4 to 

4.9, beginning early in the century. 

The estimates of the magnitude of lake 

acidification range from fractions of a pH 

unit to about 2.0, depending on the lake and 

the region. Chrysophyte analyses tend to 

infer a lower pH because they are most 

abundant in the spring and therefore reflect 

the lower pH spring condition, and because 

they may respond to episodic acidifications 

(Cumming et al. 1994). The average 

diatom-inferred pH decline in New England 

lakes is about 0.3 pH units (Davis et al. 

1994). Sullivan et al., (1990) inferred 

similar average changes for the 

Adirondacks. Cumming et al. (1994) 

examined 14 Adirondack lakes that had 

acidified since 1900 (mean pre-industrial pH 

= 5.6, range 4.9 to 6.2), and inferred a mean 

chrysophyte-inferred pH decline of about 

1.1 pH units. Kahl et al. (1989, 1991) 

observed that no clearwater (low DOC) 

lakes in Maine had pH less than 5, and 

concluded that organic acidity was 

necessary to acidify lakes to less than pH 5 

at levels of acidic deposition in the 1980s.

Figure 31. Relationship between summer and spring ANC values at LTM sites in New England, the 

Adirondacks and Northern Appalachian Plateau. Values are mean summer values for each site during the 

period 1990-2000 (horizontal axis) and mean spring minima for each site for the same time period. On 

average, spring ANC values are 30 μeq/L lower than summer values. 
62

Table 8. Estimates of change in number and proportion of acidic surface waters in acid-sensitive regions of the North and E

current rates of change in Gran ANC to past estimates of population characteristics from probability surveys. 

Region Population 

Size 

Number 

Acidic1

% 

Acidic2

Time 

Period of 

Estimate 

Current 

Rate of 

ANC 

change3

Estimated

Number 

Currently

Acidic 

Current % 

acidic 

% Ch

in Nu

of A

Syst

New England 6,834 lakes 386 lakes 5.6% 1991-94 +0.3 374 lakes 5.5% -2Adirondacks. 1830 lakes 238 lakes 13.0% 1991-94 +0.8 149 lakes 8.1% -38No. Appalachians 42,426 km 5,014 km 11.8% 1993-94 +0.7 3,600 km 8.5% -28Ridge/Blue Ridge 32,687 km 1,634 km 5.0% 1987 -0.0 1,634 km 5.0% 0% Upper Midwest 8,574 lakes 251 lakes 2.9% 1984 +1.0 80 lakes 0.9% -68

1 Number of lakes/streams with Gran ANC<0 in past probability survey (data collected at “Time Period of Estimate”, in column 2 Percent of population (from Column 2) with Gran ANC<0 in past probability survey (data collected at “Time Period of Estimat3 Based on regional trends presented in this report, in μeq/L/year 
63

Figure 32. Paleolimnological reconstruction of historical pH change in two New England LTM 

lakes, as inferred from diatom remains (from Davis et al. 1994). 

With the exception of lakes in the heavilyimpact region around the Sudbury, Ontario, 

studies of pre-industrial lake pH agree that 

the pH of presently acidified lakes was less 

than pH 6 with rare exceptions, before the 

onset of acidic deposition. Therefore, we 

propose that pH 6 is an upper limit for 

recovery of acidified lakes, equivalent to an 

Gran ANC of about 20-30 μeq/L in all 

regions. 

The interaction of S, N, and cations. 

Interactions between trends in SO4

2–, NO3–

and [Ca2+ + Mg2+] govern the magnitude of 

trends in Gran ANC. At East Bear Brook in 

Maine (Norton et al. 1999), both SO4

2– and 

NO3

–

have declined significantly since 

sampling began in 1998. If the main 

regional change in surface water chemistry 

had been a decline in SO4 without any other 

changes in major ion chemistry or DOC, 

then regional increases in ANC would have 

followed. Instead of a simple ANC 

response, there have been substantial 

declines in base cations, and modest 

increases in DOC (Figure 33). 

The relationship among these changes is 

illustrated in the response of East Bear 

Brook at BBWM (Figure 33).Concentrations 

of SO4

2– and NO3–

 have declined in parallel, 

which is presumably coincidental (Figure 

33a). The rate of decline in base cations is 

faster than the decline in SO4

2– (Figure 33b). 
64

The large decline in base cations apparently 

results from a change in soil ion exchange 

processes, or from an increase in biomass 

uptake, because there is no corresponding 

change in concentrations of silica, an 

indicator of mineral weathering rates (Figure 

33c). Silica concentrations in the LTM and 

TIME networks were also statistically 

unchanged during the 1990s. 

The trend in ANC can also be estimated 

from: 

 ANC = [Ca2+ + Mg2+ Na+

 + K+] minus 

 [SO4

2– + NO3–

 + Cl–] 

In this low DOC stream, the estimated and 

measured Gran ANC agree closely, 

suggesting that acid anion inputs and 

buffering by base cations are indeed the 

likely controlling processes for ANC. It is 

also possible to calculate the ANC change 

that would have occurred if NO3

–

 had not 

declined in 1990s (Figure 33d), by using the 

equation above, but holding NO3

–

 constant. 

The resulting calculated ANC in 1998 would 

be -20 μeq/l, instead of zero. It is possible 

that the observed regional decline in NO3

– 

during the 1990s prevented additional 

acidification in sensitive systems. 

CONCLUSIONS 

Trends in deposition and surface water 

chemistry. The CAAA has been successful 

in reducing the emissions of SO2 and 

deposition of SO4

2– accelerating the trend in 

place since about 1970. There was little 

change in NOx emissions and deposition 

until the late 1990s, when a slight decrease 

occurred due to Phase I controls on nitrogen 

emissions beginning in 1996. In response, 

SO4

2– concentrations in surface waters 

declined across all the glaciated regions of 

the North and East (Figure 34). South of the 

glaciated limit, SO4

2– concentrations have 

increased slightly, due to the re-equilibration 

of SO4

2– retaining soils with changing 

deposition. Nitrate concentrations exhibited 

few significant trends: NO3

–

 decreased 

slightly in regions with high NO3

–

, and was 

unchanged in regions with low ambient 

NO3

–

 concentrations. 

Base cations, which are important for 

neutralization of acidity in precipitation and 

in watersheds, increased non-significantly in 

atmospheric deposition during 1990-2000. 

However, the 20 year decline in base cations 

in surface waters continued, offsetting some 

of the decline in surface SO4

2–. As a result, 

Gran ANC and pH have increased less than 

might be expected from SO4

2– trends alone. 

Significant increases in Gran ANC were 

recorded in the Adirondack Mountains, the 

Northern Appalachian Plateau and in the 

Upper Midwest (Figure 34). Lakes in New 

England show little sign of recovery in the 

1990s, and streams in the Ridge/Blue Ridge 

province may even be acidifying slightly. 

Expectations for recovery. While the 

widespread declines in SO4

2– reported here 

for sensitive surface water regions 

(exclusive of the Ridge/Blue Ridge region) 

are important signs of the success of 

emission control programs, the true test of 

Title IV’s effectiveness is a reduction in the 

acidity, and the eventual biological 

recovery, of lakes and streams in the U.S. 

We report here that, in some regions, the 

number of acidic lakes and streams has 

declined significantly in the past 10 years 

(Table 8). In particular, many lakes in the 

Adirondacks and Upper Midwest, and 

streams in the Northern Appalachian 

Plateau, have achieved positive Gran ANC 

values since implementation of the CAAA. 

It is important to recognize, however, that 

despite a roughly 1/3 reduction in the 

number of acidic systems in these regions, 

8% of Adirondack lakes, and 8% of 

Northern Appalachian streams remain 

acidic. It is beyond the scope of this report 

to forecast whether continued emissions 

reductions will substantially reduce these 

numbers, or whether further additional 

regulations may be necessary. 
65

Figure 33. The interaction of SO4

2–, NO3–

 and base cations in East Bear Brook, the reference 

watershed at the Bear Brook Watershed in Maine. 
66

Uncertainties in the future behavior of NO3

–

and base cations make it very difficult to say 

how long the current recovery trends will 

continue. 

It is also important to note that one region, 

the Ridge/Blue Ridge province, has yet to 

show any signs of recovery, either in SO4

2–

concentrations or in acidity. The proportion 

of streams in this region that are acidic is 

currently less than 1%, but this region has 

the highest likelihood of undergoing further 

acidification as the capacity of regional soils 

to adsorb atmospheric SO4

2– reaches 

capacity. The small increase is surface water 

SO4

2– we report for this region should be a 

red flag to regulators that this region 

deserves further research and assessment. 

Research needs. Scientists and regulators 

have long noted the frustrations associated 

with dealing with complex environmental 

issues without long-term baseline data. With 

the continuing evolution of the chemistry of 

atmospheric deposition, data such as are 

represented in the TIME and LTM programs 

will be the only method of assessing 

recovery of surface water chemistry. LTM 

and TIME will also provide baseline sites 

for future biological assessments of 

recovery.

Figure 34. Summary of surface water trends in SO4

2–, NO3–

 , Gran ANC, base cations [Ca2+ + 

Mg2+], DOC and aluminum, by region. 
67

Importantly, these data will also prove 

useful for relevant environmental 

management issues such as base cation 

depletion, nitrogen saturation, effects of 

drought, forest health, and the myriad 

potential ecosystem-level effects of climate 

change. Increases in dissolved organic 

carbon in most lake populations may have 

contributed natural organic acidity to surface 

waters, further complicating our 

interpretation of the response to the CAAA. 

This factor is an important long-term 

research question that is not well 

understood. These LTM and TIME sites 

are core research locations that will provide 

essential information for a number of 

present and future research questions faced 

by EPA. Their long-term data serve as the 

foundation for other ecological research to 

be conducted using the locations and their 

data. Without such data, our ability to ask 

the right questions is reduced, and our 

ability to base the answers to these questions 

on actual data is seriously compromised. It 

is not possible to re-create or replace longterm data once the data record is interrupted. 

The more long-term data we accumulate, the 

better we can differentiate competing 

controls on surface water chemistry, and the 

better we can identify slow and subtle 

emerging trends, like the increased 

contribution of DOC in surface waters.
68

TERMS AND ACRONYMS USED IN THIS REPORT 

acidic describes lake or stream water with Gran ANC less than zero, corresponding to 

pH values less than ca. 5.5. 

acidified describes lake or stream water with sufficient historical information, or 

appropriate chemical relationships, to justify classification as a water that was 

historically less acidic than shown by recent measurement. The term may also 

apply to stream water which is temporarily acidified during a hydrologic episode. 

ANC Acid Neutralizing Capacity – a theoretical estimate of the ability of water to 

buffer acid (similar to ‘buffering capacity’ or ‘alkalinity’). ANC may be either 

calculated (‘calculated ANC’ = sum of cations – sum of anions) or measured 

(‘Gran ANC’) by titrating a sample of lake or stream water with a known 

concentration of acid. 

anion a negative charged ion, especially sulfate (SO4

-2) and (nitrate NO3-

) of importance 

to this report. Also see cation. 

alkalinity See ANC 

ALSC Adirondack Lake Survey Corporation 

BBWM Bear Brook Watershed in Maine. The original EPA Watershed Manipulation 

Project location for a paired-watershed experiment of enhanced S and N 

deposition (in eastern Maine) begun in 1987, with the experimental treatment 

continuing since 1989. 

CaPMon The Canadian equivalent to the U.S. NADP. 

cation A positively charged ion, especially hydrogen (H+), and the base cations calcium 

(Ca+2), magnesium (Mg+2), potassium (K+

) and sodium (Na+). See anion.

DOC Dissolved organic carbon 

CAA Federal Clean Air Act of 1970. 

CAAA Federal Clean Air Act Amendments of 1990. 

CASTNet Clean Air Status and Trends Network, the U.S. EPA-operated dry deposition 

network located primarily in the eastern U.S. 

ELS Eastern Lake Survey. The 1984 EPA program which established the baseline 

chemical condition of lakes in the eastern U.S. using a statistical survey of lake 

subpopulations. 

EMAP Environmental Monitoring and Assessment Program. The U.S. EPA monitoring 

program begun in 1991, which includes TIME and LTM. 

EMAP-SW EMAP-Surface Waters. The aquatic module within the larger environmental 

monitoring mission of EMAP. 

GAO Governmental Accounting Office 
69

Gran ANC Gran Acid Neutralizing Capacity – a measured estimate of ANC, based on using 

the Gran technique to find the inflection point in an acid-base titration of a water 

sample (Gran 1952). 

HELM High Elevation Lakes in Maine. The 90 lakes higher than 600 meters which, as a 

population, are as acidified as lakes in the Adirondack NY region. 

LTM The U.S. EPA Long-Term Monitoring project. This EPA monitoring program 

began in 1983 and included low ANC lakes in Maine, New York (Adirondacks) 

and Vermont, and low ANC streams in New York (Catskills), Pennsylvania and 

Virginia. Before 1995 it also included lakes in Colorado and the Upper Midwest 

(Wisconsin, Michigan and Minnesota). 

NADP/NTN National Atmospheric Deposition Program/National Trends Network, the 

U.S.G.S.-led entity that operates the 200+ site network of wet-only precipitation 

collectors in the U.S. 

NAPAP National Acidic Precipitation Assessment Program, the program under which 

federally-sponsored acid rain research was funded in the 1980s. NAPAP was 

responsible for the 1990 assessment report, and has continued a integrative role 

into 2002. 

NSS U.S. EPA National Stream Survey. 

NSWS U.S. EPA National Surface Water Survey. Consisting of the Eastern Lake Survey 

(see ELS), the Western Lake Survey (see WLS) and National Stream Survey (see 

NSS). 

PIRLA Paleolimnological Investigations of Recent Lake Acidification – a project 

conducted in the 1980s using diatom remains in lake sediment cores to reconstruct 

historic chemistry trends. 

recovery A return to a more pre-industrial condition, in chemistry or biology. Chemically, 

recovery would be considered an increase in alkalinity (ANC) and/or an increase 

in pH, both equivalent to a decrease in acidity in surface water. 

TIME The U.S. EPA Temporally Integrated Monitoring of Ecosystems project. TIME is 

the component of EMAP targeted at monitoring subpopulations of lakes 

(Northeast) and streams (Mid-Atlantic) that are most likely to respond to changes 

in acidic deposition. TIME and EMAP both use probability techniques to select 

sites, so that the results can be statistically inferred to represent the population of 

interest (e.g., acid sensitive lakes and/or streams in the targeted region). 

USGS U.S. Geological Survey, Department of the Interior. 

WLS Western Lake Survey. The 1986 EPA program which established the baseline 

chemical condition of lakes in the western U.S. using a statistical survey of lake 

subpopulations. 

WMP U.S. EPA Watershed Manipulation Project. See BBWM. 
70

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