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 27711ii 

## Notice

The research described in this report has b een funded by the United States Environmental Protection Agency. This document has been re viewed 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 Developmen t 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 ar e a product of the 1990 Clean Air Act Amendments (CAAA),

which utilize the successful ma rket-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 plan ts, industries, and

other sources in an effort to re duce sulfur emissions that contri bute to the development of acid

rain. EPA and collaborators have conducted ex tensive 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 Ea stern United States most affected by acid rain.

In three regions, one-quarter to one-third of lakes and stre ams previously affected by acid rain are no longer acidic. That’s good news fo r the environment and good news for the marketbased approach.

Sincerely,

Paul Gilman, Ph.D.

Assistant Administratoriv 

## Authors

John L. Stoddard 1 , Jeffrey S. Kahl 2 , Frank A. Deviney 3 , David R. DeWalle 4 , Charles T.

Driscoll 5 , Alan T. Herlihy 6 , James H. Kellogg 7 , Peter S. Murdoch 8 , James R. Webb 9 , and

Katherine E. Webster 10

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 Envi ronmental 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 Envi ronmental 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 St atistics, Colorado State University



## Acknowledgements

The quality of this report was improved significa ntly by the comments of the reviewers. The authors greatly appreciate their efforts and c onstructive input. Supplemental funding from the U.S. EPA through a USGS interagency agreemen t (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 th e 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 emi ssions following implementation of Phase I ...................21 Figure 6. Trends in wet sulfate depo sition in the eastern United States......................................21 Figure 7. Trends in wet deposition c oncentrations of sulfate during 1990-2000 ........................23 Figure 8. Trends in nitrogen oxide emi ssions following implementation of Phase I ..................24 Figure 9. Trends in wet NO 3 - deposition in the eastern United States.........................................25 Figure 10. Cumulative freque ncy distribution of slopes for sulfate and nitrogen ........................26 Figure 11. Distribution of slopes for acidity and base cations trends...........................................28 Figure 12. Annual precipita tion in the study regions....................................................................30 Figure 13. The relationship between Gran ANC and [CA 2+ + Mg 2+ ]...........................................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 i on trends in LTM network..........................................................45 Figure 23. Regional [CA 2+ + Mg 2+ ] trends in LTM network........................................................47 Figure 24. Regional DOC trends in LTM network.......................................................................48 Figure 25. Relationship between tr ends in DOC and trends in SO 4 2- ...........................................49 Figure 26. Regional aluminum trends in LTM network...............................................................50vi Figure 27. Comparison of trends in SO 4 2- in TIME probability sites ...........................................54 Figure 28. Comparison of trends in ANC in TIME probability sites ...........................................55 Figure 29. Comparison of tre nds in base cations in TIME probability sites ................................57 Figure 30. Comparison of % change in SO 4 2- concentration........................................................58 Figure 31. Relationship between summer a nd spring ANC values at LTM sites.........................61 Figure 32. Paleolimnological reconstr uction of historical pH change..........................................63 Figure 33. The interaction of SO 4 2- , NO 3 - 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 result s 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 populati ons of sites in acid sensitive regions.......................53 Table 8. Estimates of change in numb er and proportion of aci dic surface waters ......................62vii 

### 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 ch emistry in the northern and eastern U.S., in response to changes in deposition. The regions covered in this re port are New England (sites in Maine, New Hampshire, Vermont and Massachuse tts), the Adirondack Mountains of New York, the Northern Appalachian Plateau (New York, Pennsylvania and We st 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) wh ether the declines in

emissions translate into reductions in acidic de position; and (b) whether biologically relevant water chemistry has improved in acid sensitive regions. The measures of expected recovery include decreased acidity, sulfate, and t oxic dissolved aluminum concentrations.

Anthropogenic acidity in atmospheric deposition. NO

x and SO

x from the combustion of fossil fuels react with water in the atmosphere to pr oduce acid rain, a dilute solution of nitric and sulfuric acids. This acidity (and the acid anions sulfate and nitrate) ma y 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 Penns ylvania and New York, to 4.8 in Maine and the Upper Midwest. The acidity (hydrogen ion concentra tion) 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 Surf ace Water Survey (NSWS)

estimated the number of acidic waters at 4.2% oflakes and 2.7% of stream segments in acidsensitive regions of the North a nd East. Acidic waters are de fined as having acid neutralizing capacity (ANC) less than zero (i.e., no acid bufferi ng capacity in the water), corresponding to a pH of about 5.2.

This report addresses the recent ch emical responses in the surface wa ters 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 (Envir onmental Monitoring and Assessment Program). The regions include lake s in the Adirondacks, central and northern New England, and the upper Midwest. Sensitive regi ons with small streams are found in the midAtlantic region, including the nor thern and central Appalachian Pl ateau and the Ridge and Blue Ridge provinces. Surface waters in most other regions are not sensitive to the impacts of

acidification due to the na ture 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 signi ficantly at a rate between -0.75 and -1.5 μeq/L/year. There was a shar p 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 increase d 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, a nd 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 an alyses 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 surfaceix waters, with rates ra nging from -1.5 to -3 μeq/L/year (Figur e B). These declines were consistent with the decline in sulfate in pr ecipitation. Nitrate concentrations decreased in two regions with the highest ambient nitrate con centration (Adirondacks, Northern Appalachian Plateau) but were relatively unchanged in regions with low con centrations. Dissolved Organic Carbon (DOC) increased in each region, potentially contributing natural organic acidity to offset the recovery from decreased acidity a nd sulfate in deposition.

Acid neutralizing capacity is a ke y indicator of recover y, as it reflects the capacity of watersheds to buffer inputs of acidity. We expect increasin g 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 Appa lachian Plateau and Uppe r Midwest) at a rate of +1 μeq/L/year, despite a decl ine in base cations (calcium + magnesium) in each region (Figure B). The decline in base cations offsets some of th e decline in sulfate, and thus limits the increase in ANC or pH. In the Adirondacks and Northe rn Appalachians, surface water ANC and pH both increased significantly in the 1990s; toxic aluminum concentrations also dec lined 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 th an 0, or 8.1% of the population, compared to 13% (240 lakes) in the early 1990s. In the Uppe r Midwest, an estimated 80 of 250 lakes that were acidic in mid-1980s are no longer acidic. TIME surveys of st reams in the northern Appalachian Plateau region estimat ed that 5,014 kilometers of str eams (ca. 12%) were acidic in 1993-94. We estimate that 3,600 kilometers of stream s, or 8.5%, remain aci dic 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 acid ic, 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 wa ters remain relatively unchanged. There is no

evidence that the number of acidic waters have in creased in any region, de spite a general decline in base cations and a possible incr ease in natural organic acidity.

Do changes in deposition translat e 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 fo r 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 resp onse. However, the lakes and streams with the steepest declines in sulfate ha d 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 co rrespondence between rates of su lfate decline in streams and deposition in the Ridge and Blue Ridge provinces due to the adsorp tive capacity of the soils in the region. In the upper Midwest, th e 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 seepag e lakes in New England that did not experience the 1980s drought.

Complications for assessing recovery. Declines in atmospheric depos ition 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 su rface waters and marks a success of the CAAA and efforts by industry in reducing SO 2 emissions. However, the antic ipated 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 occu rred 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 a nd 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 Northe rn Appalachian Plateau); changes in nitrate are unrelated to changes in nitr ogen 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 im portant in determining the recovery, or lack of recovery, in surface waters of th e northern and eastern U.S. C ontinued long-term research and monitoring will be necessary to understand the caus es, effects, and trends in these processes.

1) Base cations . We report declining surface water c oncentrations of base cations (e.g., calcium, magnesium) in all of the glaciated re gions in this report (the Ridge and Blue Ridge region is the only non-glaciated regi on). At some indi vidual sites, further

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

While decreases in base cation loss from wate rsheds probably indica tes 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 betw een base cations and sulfate an d the effects of cation loss on soil and surface water recovery.

2) Nitrogen. Continued atmospheric loading of ni trogen may be influencing the acid-base status of watersheds in yet undetermined ways. Unlike sulfate, concentrations of nitrogen in deposition have not changed substa ntially in 20 years. Also unlike sulfate, most nitrogen deposited from the atmosphe re is retained in watershed soils and vegetation; nitrogen sequestrati on 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 char acterized by high nitrate concentrations a decade ago (Adirondacks, Northern Appal achian 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 hi story. 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 pr obably linked to complex issues including climate change and

forest maturation.

4) Climate . Climatic fluctuations induce variabil ity in surface water chemistry and thus obscure changes that we expect to result fr om 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 undersco re 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 wa tersheds 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 scal e of currently available monitoring data. Moreover, the changes observed are not unidire ctional. 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 o ccurred in several regions, additional factors appear to point toward recovery, forecasting an improvement in biologicallyxii relevant surface water chemistry. It is not ye t clear if further reductions in emissions and deposition will be necessary for widespread recove ry to occur. These factors forecast the onset of recovery:

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

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

c) Dissolved Organic Carbon has increased re gionally, perhaps towarda 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 acidifie d 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 recove ry 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 eviden ce of recovery from the effects of acidic deposition. The comple xities of ecosystem response – eff ects of forest health, soil status, natural organic acidity, the re lative 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 tr end 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 chem istry 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 th e 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 watershe ds and surface waters to a host of research and regulatory issues related to acidic deposition, including soil a nd surface water recovery, controls on nitrogen retention, mechanisms of base cation depletion, fo rest health, sinks for sulfur in watersheds, changes in DOC and speciation of aluminum, and various factors re lated to climate change. As one reviewer of this report noted, ”...these sites have irreplaceabl e long-term data that should constitute a ‘research infrastructure’ akin to an EPA laboratory. These sites will help address many basic science issues in whic h EPA ORD has a continuing interest....” Moreover, asxiii several of the reviewer s 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 ques tions 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), withidentification 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 SO 2 and NO

x , 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 Uppe r 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 SO 2 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.

NO X 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 change s 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 SO 2 , and 2 million ton reduction in NOx, by 2010, compared to the baseline year of 1980. Emissions of SO 2 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 SO 2 — 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 SO 2 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 SO 2 trading program, with flexibility in the method to achieve emission reductions. However, it does not cap NOx

emissions as the SO 2 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 reduc tions 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 th e 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 de position 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 re gional 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 therefor e 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., Wi gington 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 etal. 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 da tasets 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 sites 1 Size of Population 2 Percent acidic in

1980s 3

Statistical Surveys

New England Lakes 4 30 4,327 lakes 5%

Adirondack Lakes 4 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 fo r this report (1990-2000); locations of sites are illustrated in Figures 2 and 3.

2 Total number of lakes, or stream length, for whic h statistical survey results can be inferred. Site selection for LTM (sensitive surface waters) is not sta tistically based, and results cannot be expanded to population level.

3 Estimates of extent of acidification, based on Natio nal 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 popula tions 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 7.6Figure 2. Location of acid-sensitive regions of the northern and eastern U.S. covered in this report, and locations of individual LTMsites 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 c oncentrations. Higher dissolved organic carbon (DOC) concentrations have been reported ubiquitously in surface waters of the northern hemisphere in the past decade (Evans and Monteith 20 01, Skjelkvåle et al. 2001, Jeffries et al. 2002; this report). Higher DOC may be an indicator of recovery, due to a dec line 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 discusse d 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 widespre ad 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 (NH 3 ) 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 (NH 4 + ) 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,

SO 2 reacts with water in the atmosphere to yield sulfuric acid:

SO 2 + H 2 O + 1⁄2O 2 Æ H 2 SO 4 An analogous reaction of water with nitrogen oxides, symbolized as NO

x , yields nitric acid (HNO 3 ).

In addition to wet deposition (rain, snow, and fog), acidic deposition includes the deposition of dry, part iculate, 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 rainfa ll (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 hydr ochloric 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 th at 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 precipitati on is still subject to misunderstanding. Even in pristine environments, precipitation pH is rarely controlled by the CO 2 reaction that has an equilibrium pH of 5.6:

H 2 O + CO 2 Æ H 2 CO 3 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 Precipitation9 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 po tential 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 inte rpret 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 concentratio ns 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 acidifica tion 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 anthr opogenically 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 regions10 translate to several hundred acidic waters in each region.

The Valley and Ridge province and Northern Appalachian Plateau had 5% and6% acidic sites, respectively. The only acidsensitive region not assessed in the current report is Florida, where the high proportion of naturally acidic lake s, 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 de leterious to biota as chronic acidification (B aker et al. 1996). The mechanisms that produce acidic episodes include dilution, and flushing of

NO 3 - , SO 4 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

NO 3 - during high flow events), producing hydrologically-mobile stores of SO 4 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, al though 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 bi oassays (e.g., Baker et al. 1996), with whole-ecosystem acidification experime nts (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 th e effect of acidic deposition on the distribution of sensitive11 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 de position 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 th e 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 at tention 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 NO 3 -

. Nitrate export can cont ribute to acidification (especially episodic acidification), mobilization of aluminum, and leaching of cations from soils (A ber 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 concen trations 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 NO 3 – concentrations, particul arly 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 (NO

x ) and agricultural sources (NH 3 and NO

x ), 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 fu ture 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 th e regional boundaries13 slightly to include deposition monitoring

stations within clos e 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 concentratio ns of ions in surface waters with the concentrations in deposition. At present there is no straightforward method to produce deposition concentrations (e.g., SO 4 2– in μeq/L) from dry deposition data. In addition, th e 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 Fernberg15 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 Ke ntucky 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 to16 changes in rates of acidic deposition (i.e., those with Gran ANC < 100 μeq/L). In the Mid-Atlantic, the targ et population is upland

streams with a hi gh 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 Adirondack s and New England.

Data from 43 Adirondack lakes can be extrapolated to the ta rget 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-Atlan tic 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 anal ysis; 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 l ong-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 at tributable 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 hist orical 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 values17 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 reporte d 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 for18 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 lake s 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 re search 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 anal yses of relationships

between SO 4 2– and base cation behavior required the greater precision provided by19 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 an d base cations based on the ratios found in seawater (Riley and Chester 1971).

Table 3 . Typical analytical methods used for samples repor ted here. Samples are filtered before analysis, except in Adirondack LTM.

Variable Technique Protocol reference

pH, closed cell Electrode Hillman, et al 6 , EPA 19.0 5

pH, aerated Electrode Hillman, et al 6 , EPA 5.0 5

Specific conductance Wheatstone bridge EPA 120.1 2 ,EPA 23.0 5

True color Spectrophotometer, 457.5 nm EPA 110.2 2

Gran ANC Gran Titration Hillman et al 6 , EPA 5.0 5

Anions: Cl

- , NO 3 - , SO 4 2- Ion chromatography EPA 300.0 1

Calcium AAS with N 2 0-acetylene flame EPA 215.1 2

Magnesium AAS with N 2 0-acetylene flame EPA 242.1 2

Sodium AAS with air-acetylene flame EPA 258.1 2

Potassium AAS with air-acetylene flame EPA 273.1 2

Aluminum, total AAS with graphite furnace EPA 200.9 4

Dissolved Organic Carbon IR C analyzer, persulfate oxidation EPA 415.1 2

Dissolved Inorganic Carbon IR C analyzer, direct injection EPA 13.0 5 , OI, 1990 8

Ammonium Autoanalyzer EPA 9.0 5 and Bran & Luebbe 780-

86T 7

Silica Autoanalyzer EPA 22.0 5 and Bran & Luebbe 785-

86T 7

AAS=atomic absorption spectrophotometry IR=Infrared Spectrophotometry

Method references:

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

1993.

2 Methods for Chemical Analysis of Water a nd 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. Anal ytical 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, 199020 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 SO 2 (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 east ern United States from NADP/NTN data (1989-1991 vs. 1995-1998). Data are in kg/ha.22 Power plant SO 2 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 SO 4 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. SO 2 emissions have decreased by about 5.5 million tons. Decreases in NO

x emissions were more modest than those of sulfur, with declines

noted in several states in the northeastern corridor (Figure 6). Si nce 1990, total utility

NO

x 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 onl y about one-third of

total NO

x emissions. Since total NO

x emissions from other sources have remained relatively constant (mot or 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 NO

x emissions.23 Figure 7 . Trends in wet deposition concentrations of SO 4 2- during 1990 – 2000 (NADP, 2002). Boxes

indicate the interquartile range in SO 4 2- for each region in each year; caps indicate the 10 th and 90 th
percentiles; median concentration is indicated by line in box; dots enclose the range.24 Figure 8 . Trends in nitrogen oxide emis sions following implementation of Phase I of the Acid Rain Program: Total state-level utility NO

x (1990 vs. 1999).

Relatively constant NO

x emissions translate to little change in wet NO 3 – deposition in the 1990s (Figure 8). The relative lack of change in NO 3 – deposition, relative to the

decline in SO 4 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 SO 4 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 precipita tion 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, a nd for which statistical significance is calculat ed, can be estimated by finding the 50 th percentile value for the region (from the y-axis) and estimating the corresponding value on the x-axis. For example, the median change in SO 4 2–
concentration in wet deposition in the Northern Appalachia n Plateau region (Figure 10a) is ca. -1.6 μeq/L/year.25 Figure 9 . Trends in wet NO 3 - deposition in the eastern United Stat es 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 (NO 3 - plus NH 4 + ) 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 SO 4 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 SO 4 2- by reducing sulfur emissions. Wet deposition of SO 4 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 NO 3 - concentrations also declined significantly in every eastern region, with the exceptio n of the Ridge and Blue Ridge provinces, where declines were not significant. These tr ends appear to be driven by a decline in NO 3 - in deposition after 1995. Nitrogen and NO 3 - 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 th at 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 SO 4 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 SO 4 2- (panel a, top) and nitrogen (panel b, bottom) trends in wet deposition, from regiona l 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 regi on corresponds to the value of the x-axis at the

50 th percentile of the curve.27 Table 4 . Regional trend results (1990-2000) for atmos pheric deposition (wet-onl y 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 SO 4 2- Nitrogen NO 3 - Base Cations Hydrogen

N ew England -0.96** -0.26* -0.20** +0.02 ns -0.81**

Adirondacks -1.47** -0.37** -0.38** +0.01 ns -1.48** Appalachian Plateau -1.55** -0.41** -0.37** +0.01 ns -1.54** Ridge/Blue Ridge Province -0.96** -0.16 ns -0.13 ns +0.05 ns -0.94 ns
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 (Ca 2+ plus Mg 2+ ) 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 unders tood 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

SO 4 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

SO 4 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 (Ca 2+ +Mg 2+ ) (panel b, bottom) trends in regional NADP/N TN data, by region, for the period 1990-2000.29 What is ‘background’ precipitation chemistry? Miller (1999) analyzed the relationship between NADP/NTN SO 4 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 SO 4 2- deposition to a hypothetical condition of zero SO 2 emissions. The predicted range of precipitation SO 4 2- concentrations at zero

SO 2 emission was 1.9 to 4.4 kg/ha at the six stations for which Miller developed equations. The mean SO 4 2- deposition intercept (i.e. at ze ro emissions), was 2.9 kg/ha. Assuming consistent climatic conditions, this deposition would equate to 6 to 10 μeq/L SO 4 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 SO 4 2- at
NADP/NTN sites in the eastern U.S., and 5 to 20 μeq/L in the western U.S. The precipitation pH re presented by this background SO 4 2- would be pH 5.0 to 5.2, not including acidity fr om nitrate or organic acids. This analysis supports the inference ofCharlson 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 unlikel y 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 mayhave the effect of isolating lakes from baserich groundwater, es sentially 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 (F igure 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 extr a 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 tr ends in the east and northeast with one excepti on: recovery from the drought of the 1980s in the Upper Midwest (Webster and Brezonik 1995) is

believed to have influe nced 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) SO 4 2

- and NO 3 - , 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 SO 4 2– ) and in ecosystem response to long-term deposition (e.g., NO 3 – ).30 Figure 12 . Annual precipitation in the study regions (N ational 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 [Ca 2+ + Mg 2+ ], 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 [Ca 2+ + Mg 2+ ]) 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 [Ca 2+ + Mg 2+ ] 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,

SO 4 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 th is 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.

S ampling 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 Upper33 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 SO 4 2– , NO 3 – , base cations [Ca 2+ + Mg 2+ ], Gran ANC, pH, and DOC in Salmon Pond, Maine (New England region). Signif icant trends are indicated by trend lines. Shaded box indicates time peri od of analyses reported here.35 Figure 15 . Time series data for SO 4 2– , NO 3 – , base cations [Ca 2+ + Mg 2+ ], Gran ANC, pH, and DOC in Darts Lake, NY (Adirondack region). Significant tr ends are indicated by trend lines. Shaded box indicates time period of analyses reported here.36 Figure 16 . Time series data for SO 4 2– , NO 3 – , base cations [Ca 2+ + Mg 2+ ], 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 SO 4 2– , NO 3 – , base cations [Ca 2+ + Mg 2+ ], 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 SO 4 2– , NO 3 – , base cations [Ca 2+ + Mg 2+ ], 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. SO 4 2– declined

in surface waters in th e 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 exce ption to the pattern of declining SO 4 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 SO 4 2– increased slightly in Ridge and Blue Ridge streams. All of the regional SO 4 2– trends are highly significant (Table 5), and are consistent with the trends reported previously, in cluding other regions of North America and Europe (Mattson et al. 1997, Stoddard et al. 1999, Evans and Monteith 2001, Skjelkvå le et al. 2001).

The marked declines in SO 4 2– concentrations in the glaciated portions of the North and East are almost certainly direct responses to declining emissions and SO 4 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 SO 4 2– that has been occurring for almost 3 decades (Stoddard et al. 1999).

The small increase in SO 4 2– concentrations in the Ridge and Blue Ridge provinces

results from the SO 4 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 SO 4 2– well below those found in deposition (Figure 7). The future response of

SO 4 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 [Ca 2+ + Mg 2+ ], Gran ANC and hydrogen are μeq/L/year. Units for DOC are mg/L/year. Units for aluminum are μg/L/year.

Region SO 4 2– NO 3 – Base

Cations Gran

ANC Hydrogen DOC Aluminum

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

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* +0.03 ns +0.56 ns

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

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

ns regional trend not significant (p > 0.05) * p < 0.05 ** p < 0.01

NA insufficient data40 Figure 19 . Cumulative frequency diagram (distribution) of slopes (μeq/L/year) for SO 4 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 NO 3 –

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

Trends in nitrate by region. In general,

trends in NO 3 – were much smaller than

trends in SO 4 2– , with the only ecologically significant changes o ccurring in the two regions with the highest ambient NO 3 – concentrations (Figure 20). Both lakes in

the Adirondacks and streams in the Northern Appalachian Plateau exhibited small but significant downward trends in NO 3 – 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

NO 3 – concentrations in these regions are a positive development for these ecosystems, we clearly do not know if these trends will continue, especially si nce they do not reflect recent trends in emissions or deposition (Table 4, Figures 7-9). The presence of
strong upward trends in NO 3 – 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 (variabi lity) in a long-term pattern of changing NO 3 – leakage from42 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 cons idered (e.g., centuries rather than decades), the long-term retention nitrogen deposition in forested regions is unlikely to continue inde finitely (Aber et al. in press).

In New England and the Upper Midwest,

where ambient NO 3 – concentrations are

much lower than in the Adirondacks and Northern Appalachian Plateau (Figures 14- 17), NO 3 – concentrations in surface waters were unchanged. The Ridge/Blue Ridge province registered a sm all, but significant,

decrease in NO 3 – during the 1990s, but interpretation of trends for NO 3 – 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 NO 3 – exportfrom 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 wo uld 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 th e 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 highli ghted 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 patt ern shown for Darts Lake in Figure 15 and the Neversink River in Figure 16, with Gran ANC increasingafter 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 th e 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 exhi bited by lakes in the Adirondacks, with long-term recovery in Gran ANC having fallen off after a similar inflection point in th e 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 SO 4 2– observed. It is cl ear that, while the evidence for a widespread recovery in ANC in the glaciated North and East is quite43 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 [Ca 2+ + Mg 2+ ] 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 (Ga lloway 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 Ca 2+ has been occurring for more than a century in Ontario lakes.

As a broad generalization, SO 4 2– has decreased at a rate of approximately –2.5 μeq/L/year (the mean of regional median slopes), and NO 3 – 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 SO 4 2– and NO 3 – 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 [Ca 2+ + Mg 2+ ] was about –2.0 μeq/L/year.44 Figure 21 . Cumulative frequency diagram (distributi on) 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 [Ca 2+ + Mg 2+ ], while not unexpected, limits the magnitude of ANC recovery by offsetting so me of the decline in

SO 4 2– and NO 3 –. . 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 analys es (Stoddard et al. 1998a, Stoddard et al. 1998b) using data through the early 1990s. The observation that acid anions, and particularly SO 4 2–

, concentrations are decrea sing 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 th e 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 cr itical levels. At least one aspect of the current analysis argues against this hypothesis. The regional response of base cations to changing SO 4 2–
deposition is following expectations (e.g., [Ca 2+ + Mg 2+ ] is decreasing at lower rates

than SO 4 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 wate rsheds 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 th e 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 (Skjelkvål e et al. 2001). This argues for a large-scale cause, and both climate change (warmer temperatures are expected to increase decomposition, and could potentially increas e 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., SO 4 2–

and NO 3 – ) 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 reconstr uctions 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). Davi d et al. (1999) were also able to measure a decrease in organic anions in response to the experimental

whole-watershed acidifi cation experiment at

the Bear Brook Waters hed in Maine (Norton et al. 1999). In one of the few studies of DOC behavior during recovery from47 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 [Ca 2+ + Mg 2+ ] 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 s ites, 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 SO 4 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

SO 4 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 SO 4 2– concentrations for lakes. This graph includes data from the Maine high elevation and seep age 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 biologi cally-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 acidificati on, 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 SO 4 2– and/or

NO 3 – 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, lo w ANC and moderate ANC sites for the period 1990- 2000. Analysis includes all sites in New Engla nd, 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 se lection utilized for LTM varied among regions, and was in part driven by the existence of historical data and/or ongoing data co llection, rather than

statistical frameworks to determine population trends. Stodd ard 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 th e analysis of LTM representativeness into the realm of trends, as well as changing popu lation 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 SO 4 2– to those for LTM sites (Figure 27). Both datasets suggest highly significant regional

decreases in SO 4 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 slight53 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 regi onal target populations. Lake trends in the Adirondacks and New England are for the period 19 91-2001. 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 SO 4 2– , NO 3 – , base cations [Ca 2+ + Mg 2+ ], 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 SO 4 2– NO 3 – Base

Cations Gran

ANC Hydrogen DOC Aluminum

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

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

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

ns regional trend not significant (p > 0.05) * p < 0.05 ** p < 0.0154 Figure 27 . Comparison of trends in SO 4 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 [Ca 2+ + Mg 2+ ]
concentrations in th e 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, SO 4 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 SO 4 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 r ecovery response for

SO 4 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 SO 4 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 SO 4 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

SO 4 2– concentrations in surface waters than in deposition. In contrast, the percent change in SO 4 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 de position 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

%SO 4 2– ) are generally steeper than in surface waters, suggesting that most aquatic systems are exhibiting a lagged response. Apparently watershed soils retain enough atmospheric SO 4 2– to reduce the rate declines occurring in deposition (Figure 30). Interestingly, the lakes and streams with the steepest declines in %SO 4 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 corresponden ce between rates of

%SO 4 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 [Ca 2+ + Mg 2+ ] in TIME probability sites (expressed as the distribution of slopes in the ta rget population) and LTM sites in New England

and the Adirondacks.58 Figure 30 . Comparison of % change in SO 4 2– concentration in wet deposition and surface waters, by region.59 In the upper Midwest, th e rate of decline in

lakes is greater than the decline in deposition, probably re flecting the residual effects of the drought of the late 1980s. All

of the lakes with str ong negative trends in

%SO 4 2– are seepage lake s 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 SO 4 2– may also be

stored in the sediments of these lakes via sulfate reduction (B aker 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. Ther e is reason to expect that additional reductions in deposition

would result in additiona l 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 syst ems 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 in dicated 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 of60 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 estimat ed 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 incr ease 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 magnit ude 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 fr om 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. Cummi ng 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. Valu es are mean summer values for each site during the

period 1990-2000 (horizontal axis) and mean spring mini ma for each site for th e same time period. On average, spring ANC values are 30 μeq/L lower than summer values.62Table 8. Estimates of change in number and proportion of acidic surface waters in acid-sensitive regions of the North and East, basedon applyingcurrent rates of change in Gran ANC to past estimates of population characteristics from probability surveys.RegionPopulationSizeNumberAcidic1%Acidic2TimePeriod ofEstimateCurrentRate ofANCchange3EstimatedNumberCurrentlyAcidicCurrent %acidic% Changein Numberof AcidicSystemsNew England 6,834 lakes 386 lakes 5.6% 1991-94 +0.3 374 lakes 5.5% -2%Adirondacks. 1830 lakes 238 lakes 13.0% 1991-94 +0.8 149 lakes 8.1% -38%No. Appalachians 42,426 km 5,014 km 11.8% 1993-94 +0.7 3,600 km 8.5% -28%Ridge/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%1Number of lakes/streams with Gran ANC<0in past probability survey (data collected at “Time Period of Estimate”, in column 5)2Percent of population (from Column 2) with Gran ANC<0 in past probability survey (data collected at “Time Period of Estimate”,in column 5)3Based on regional trends presented in this report, in μeq/L/year63 Figure 32 . Paleolimnological reconstruction of histor ical 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-industria l 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 SO 4 2– , NO 3 – and [Ca 2+ + Mg 2+ ] govern the magnitude of

trends in Gran ANC. At East Bear Brook in Maine (Norton et al. 1999), both SO 4 2– and

NO 3 – have declined significantly since sampling began in 1998. If the main regional change in surface water chemistry

had been a decline in SO 4 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 SO 4 2– and NO 3 – have declined in parallel, which is presumably coincidental (Figure 33a). The rate of decline in base cations is

faster than the decline in SO 4 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 = [Ca 2+ + Mg 2+ Na + + K + ] minus [SO 4 2– + NO 3 – + 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 NO 3 – had not declined in 1990s (Figure 33d), by using the equation above, but holding NO 3 – constant. The resulting calculate d ANC in 1998 would be -20 μeq/l, instead of zero. It is possible that the observed regional decline in NO 3 – 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 SO 2 and deposition of SO 4 2– accelerating the trend in place since about 1970. There was little change in NO

x 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,

SO 4 2– concentrations in surface waters declined across all the glaciated regions of the North and East (Figure 34). South of the glaciated limit, SO 4 2– concentrations have increased slightly, due to the re-equilibration

of SO 4 2– retaining soils with changing deposition. Nitrate conc entrations exhibited few significant trends: NO 3 – decreased slightly in regions with high NO 3 – , and was unchanged in regions with low ambient

NO 3 – concentrations.

Base cations, which are important for neutralization of acidity in precipitation and in watersheds, increase d 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 SO 4 2– . As a result, Gran ANC and pH have increased less than might be expected from SO 4 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 SO 4 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 SO 4 2– , NO 3 – and base cations in East Bear Brook, the reference

watershed at the Bear Br ook Watershed in Maine.66 Uncertainties in the future behavior of NO 3 – 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 SO 4 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 SO 4 2– reaches capacity. The small increase is surface water

SO 4 2– we report for this region should be a red flag to regulator s that this region

deserves further research and assessment.

Research needs. Scientists and regulators have long noted the frus trations 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 SO 4 2– , NO 3 – , Gran ANC, base cations [Ca 2+ + Mg 2+ ], 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 orga nic 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 locat ions 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 ecol ogical 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-c reate 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 ju stify classification as a water that was historically less acidic than shown by r ecent measurement. The term may also apply to stream water which is tempor arily 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 ca tions – 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 i on, especially sulfate (SO 4 -2 ) and (nitrate NO 3 - ) of importance to this report. Also see cation

. alkalinity See ANC

ALSC Adirondack Lake Survey Corporation

BBWM Bear Brook Watershed in Maine. Th e 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 Netw ork, the U.S. EPA-operated dry deposition network located primarily in the eastern U.S.

ELS Eastern Lake Survey. The 1984 EPA pr ogram which establ ished the baseline

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

EMAP Environmental Monitoring and Asse ssment 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 Office69 Gran ANC Gran Acid Neutralizing Capacity – a measured estimate of ANC, based on using the Gran technique to find the inflection poi nt 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 la kes 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 lake s in Maine, New York (Adirondacks) and Vermont, and low ANC streams in Ne w York (Catskills), Pennsylvania and Virginia. Before 1995 it also included la kes 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 2 00+ site network of wet-only precipitation

collectors in the U.S.

NAPAP National Acidic Precipitation Asse ssment Program, the program under which federally-sponsored acid rain resear ch 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 Surve y. 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 ch emistry 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 bot h 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 pr ogram which estab lished the baseline

chemical condition of lakes in the western U.S. using a statistic al survey of lake subpopulations.

WMP U.S. EPA Watershed Mani pulation Project. See BBWM.70 REFERENCES

Aber, J., W. McDowell, K. Nadelhoffer, A. Ma gill, G. Berntson, M. Kamakea, S. McNulty, W. Currie, L. Rustad, and I. Fernandez. 1998. N itrogen saturation in temperate forest ecosystems. BioScience 48 :921-934. Aber, J. D., and C. T. Driscoll. 1997. Effects of land use, climate variation and N deposition on N cycling and C storage in northern hardw ood forests. Global Biogeochemical Cycles

11 :639-648. Aber, J. D., C. L. Goodale, S. V. Ollinger, M. -L. Smith, A. H. Magil, M. E. Martin, R. A. Hallett, and J. L. Stoddard. in press. Is nitr ogen deposition altering the nitrogen status of

Northeastern forests? BioScience. Aber, J. D., A. Magill, S. G. McNulty, R. D. Boone, K. J. Nadelhoffer, M. Downs, and R. Hallett. 1995. Forest biogeochemistry and primary production a ltered by nitrogen

saturation. Water Air and Soil Pollution 85 :1665-1670. Aber, J. D., K. J. Nadelhoffer, P. Steudler, and J. M. Melillo. 1989. N itrogen saturation in northern forest ecosystems. BioScience 39 :378-386. Altman, D., T. Bryant, M. Gardner, and D. Machin. 2000. Statistics with Confidence. BMJ Books, London. Baker, J. P., and S. W. Christensen. 1991. Effect s of acidification on biol ogical communities in aquatic ecosystems. Pages 83-106 in D. F. Charles, editor. Acidic Deposition and Aquatic Ecosystems. Regional Case Studi es. Springer-Verlag, New York. Baker, J. P., and C. L. Schofield. 1982. Aluminum toxicity to fish in acidic waters. Water Air

and Soil Pollution 18 :289-309. Baker, J. P., J. Van Sickle, C. J. Gagen, B. P. Ba ldigo, D. W. Bath, R. F. Carline, D. R. DeWalle, W. A. Kretser, P. S. Murdoch, W. E. Shar pe, H. A. Simonin, and P. J. Wigington. 1996. Episodic acidification of small streams in the northeastern United States: Effects on fish populations. Ecological Applications 6 :422-437. Baker, L. A., P. L. Brezonik, and C. D. Pollma n. 1986. Model of internal al kalinity generation: Sulfate retention component. Water Air and Soil Pollution 31 :89-94. Baker, L. A., A. T. Herlihy, P. R. Kaufmann, an d J. M. Eilers. 1991. Acidic lakes and streams in

the United States: The role of acidic deposition. Science:1151-1154. Baker, L. A., P. R. Kaufmann, A. T. Herlihy, an d J. M. Eilers. 1990. Current Status of Surface Water Acid-Base Chemistry. NAPAP St ate-of-Science/Technology Report No. 9, National Acid Precipitation Asse ssment Program, Washington, DC. Charles, D. F., M. W. Binford, E. J. Furlong, R. A. Hites, M. J. Mitchell, S. A. Norton, F. Oldfield, M. J. Paterson, J. P. Smol, A. J. Uu tala, J. R. White, D. R. Whitehead, and R. J. Wise. 1990. Paleoecological inves tigation of recent lake acid ification in the Adirondack Mountains, New York. Jour nal of Paleolimnology 3 :195-241. Charles, D. F., and S. A. Norton. 1986. Paleolimno logical evidence for trends in atmospheric deposition of acids and metals. Pages 335-431 in Acid Deposition: Long-term Trends. National Academy Press, Washington, D.C. Charles, D. R., editor. 1991. Acidic Deposition a nd Aquatic Ecosystems. Regional Case Studies. Springer-Verlag, New York. Charlson, R. J., and H. Rodhe. 1982. Factors controll ing the acidity of natural rainwater. Nature

295 :683-685.71 Chen, C. W., S. A. Gherini, R. J. M. Hudson, J. D. Dean, and R. A. Goldstein. 1984. Development and calibration of the Integr ated Lake-Watershed Acidification Study model. Pages 175-203 in J. L. Schnoor, editor. Modeling of Total Acid Precipitation Impacts. Butterworth Publishers, Stoneham, MA. Christophersen, N., and R. F. Wright. 1981. Sulfat e flux and a model for sulfate concentrations in streamwater at Birkenes, a small forested catchment in southernmost Norway. Water

Resources Research 17 :377-389. Church, M. R. 1999. The Bear Brook Watershed Manipulation Project: Watershed science in a policy perspective. in S. A. Norton and I. J. Fernandez, editors. The Bear Brook

Watershed in Maine: A Paired Watershed Experiment -- The First Decade (1987-1997).

Kluwer Academic Publishers. Church, M. R., P. W. Shaffer, K. N. Eshleman, and B. P. Rochelle. 1990. Potential effects of sulphur deposition on stream chemistry in th e Southern Blue Ridge Province. Water Air

and Soil Pollution 50 :39-48. Church, M. R., K. W. Thornton, P. W. Shaffer, D. L. Stevens, B. P. Rochelle, G. R. Holdren, M. G. Johnson, J. L. Lee, R. S. Turner, D. L. Cassell, D. A. Lammers, W. G. Campbell, C. I. Liff, C. C. Brandt, L. H. Liegel, G. D. Bis hop, D. C. Mortenson, S. M. Pierson, and D. D. Schomoyer. 1989. Direct/Delayed Response Projec t: Future Effects of Long-Term Sulfur Deposition on Surface Water Chemistry in th e Northeast and Southern Blue Ridge Province. Volume III. Level III Analyses and Summary of Results. U.S. Environmental Protection Agency, Washington, D.C. Cleveland, L., E. Little, S. Hamilton, D. Buckle r, and J. Hunn. 1986. Interactive toxicity of aluminum and acidity to early life stages of brook trout. Tr ansactions of the American Fisheries Society 115 :610-620. Cosby, B. J., G. M. Hornberger, R. F. Wright, and J. N. Galloway. 1986. Modelling the effects of acid deposition: Control of long-term sulf ate dynamics by soil sulfate adsorption. Water

Resources Research 22 :1283-1291. Cosby, B. J., R. F. Wright, G. M. Hornberger , and J. N. Galloway. 1985. Modeling the effects of acid deposition: Assessment of a lumped pa rameter model of soil water and streamwater chemistry. Water Resources Research 21 :51-63. Cumming, B. F., K. A. Davey, J. P. Smol, a nd H. J. B. Birks. 1994. When did acid-sensitive Adirondack lakes (New York, USA) begin to acidify and are they still acidifying?

Canadian Journal of Fish eries and Aquatic Science 51 :1550-1568. David, M., G. Vance, and J. Kahl. 1999. Chemis try of dissolved organic carbon at Bear Brook Watershed, Maine: stream water response to (NH 4 ) 2 SO 4 . Environmental Monitoring and

Assessment 55 :149-163. Davis, R. B., and D. S. Anderson. 1985. Method s of pH calibration of sedimentary diatom remains for reconstructing history of pH in lakes. Hydrobiologia 120 :69-87. Davis, R. B., D. S. Anderson, and F. Berg e. 1985. Palaeolimnological evidence that lake acidification is accompanied by lo ss of organic matter. Nature 316 :436-438. Davis, R. B., D. S. Anderson, S. A. Norton, P. R. Sweets, J. Ford, and J. S. Kahl. 1994. Sedimented diatoms in northern New England lakes and their use as pH and alkalinity indicators. Canadian Journal of Fisheries and Aquatic Science 51 :1855-1876. DeWalle, D. R., and B. R. Swistock. 1994. Causes of episodic acidification in five Pennsylvania streams on the northern Appalachian Plateau. Water Resources Research 30 :1955-1963.72 Dillon, P. J., L. A. Molot, and M. Futter. 1997. A note on the effects of El Niño-related drought on the recovery of acidified lakes. Interna tional Journal of Envir onmental Monitoring and

Assessment 46 :105-111. Dixit, S. S., W. Keller, A. Dixit, and J. P. Smol. 2001. Diatom-inferred dissolved organic carbon reconstructions provide assessments of past UV-B penetration in Canadian Shield lakes.

Canadian Journal of Fisher ies and Aquatic Sciences 58 :543-550. Driscoll, C. T., G. B. Lawrence, A. J. Bulger, T. J. Butler, C. S. Cronan, C. Eager, K. F. Lambert, G. E. Likens, J. L. Stoddard, and K. C. Weathers. 2001. Acidic deposition in the

northeastern United States: S ources and inputs, ecosyste m effects and management strategies. BioScience 51 :180-198. Driscoll, C. T., G. E. Likens, L. O. Hedin, J. S. Eaton, and F. H. Bormann. 1989. Changes in the chemistry of surface waters: 25-year result s at the Hubbard Brook Experimental Forest. Environmental Science and Technology 23 :137-143. Driscoll, C. T., R. M. Newton, C. P. Gubala , J. P. Baker, and S. W. Christensen. 1991. Adirondack Mountains. Pages 133-202 in D. F. Charles, editor. Acidic Deposition and Aquatic Ecosystems: Regional Case St udies. Springer-Verlag, New York, NY. Driscoll, C. T., K. M. Postek, W. Kretser, a nd D. J. Raynal. 1995. Long-term trends in the chemistry of precipitation and lake water in the Adirondack region of New York, USA.

Water Air and Soil Pollution 85 :583-588. Driscoll, C. T., and R. Van Dreason. 1993. Seasona l and long-term temporal patterns in the chemistry of Adirondack lakes. Water Air and Soil Pollution 67 :319-344. Eshleman, K. N., R. P. Morgan II, J. R. Webb, F. A. Deviney, and J. N. Galloway. 1998. Temporal patterns of nitrogen leakage from mid-Applachian forested watersheds: role of

insect defoliation. Water Resources Research 34 :2005-2016. Evans, C. D., and D. T. Monteith. 2001. Chemical trends at lakes and streams in the UK Acid Waters Monitoring Network, 1988-2000: Evidence for recent recovery at a national scale. Hydrology and Earth Systems Sciences 5 :351-366. Fenn, M. E., M. A. Poth, J. D. Aber, J. S. Bar on, B. T. Bormann, D. W. Johnson, A. D. Lemly, S. G. McNulty, D. F. Ryan, and R. Stottlem yer. 1998. Nitrogen excess in North American ecosystems: predisposing factors, ecosystem responses, and management strategies. Ecological Applications 8 :706-733. Ford, J., J. L. Stoddard, and C. F. Powers. 1993. Pe rspectives in environm ental monitoring: an introduction to the U.S. EPA Long-Term M onitoring (LTM) project. Water Air and Soil

Pollution 67 :247-255. Gallagher, J., and J. Baker. 1990. Current status of fish commuities in Adirondack lakes. Pages

3-11 to 13-48 in Adirondack Lakes Survey: An In terpretative Analysis of Fish Communities and Water Chemistry, 1984-87. Adir ondack Lake Survey Corporation, Ray Brook, NY. Galloway, J. N., S. A. Norton, and M. R. Church. 1983. Freshwater acidification from atmospheric deposition of sulfuric acid: A conceptual model. Environmental Science and Technology 17 :541-545. Gherini, S., R. Munson, E. Altwicker, R. April, C. Chen, N. Clesceri, C. Cronan, C. Driscoll, A. Johannes, R. Newton, N. Peters, and C. Schofield. 1989. Regional Integrated LakeWatershed Acidification Study (RILWAS): Su mmary of Major Findings. Electric Power Research Institute, Palo Alto, CA.73 Goodale, C. L., J. D. Aber, and W. H. McDowe ll. 2000. The long-term effects of disturbance on organic and inorganic nitr ogen export in the White Mountains, New Hampshire. Ecosystems 3 :433-450. Government Accounting Office. 2000. Acid Rain: Emission Trends and Effects in the Eastern United States. GAO/RCED-00-47, Govermen t Accounting Office, Washington, DC. Gran, G. 1952. Determination of the equivalence poi nt in potentiometric titrations. ANALYST: International Congress on Analytical Chemistry 77 :661-671. Hedin, L. O., L. Granat, G. E. Likens, T. A. Buishand, J. N. Galloway, T. J. Butler, and H. Rodhe. 1994. Steep declines in atmospheric base cations in regions of Europe and North

America. Nature 367 :351-354. Henriksen, A., and M. Grande. 2002. Lake Langtjern -- fish studies in the Langtjern area 1966- 2000. Report SNO 4537-2002, Norwegian Institute for Water Research, Oslo. Herlihy, A. T., D. P. Larsen, S. G. Paulse n, N. S. Urquhart, and B. J. Rosenbaum. 2000. Designing a spatially balanced, randomised site selection process for regional stream surveys: The EMAP Mid-Atlantic Pilo t Study. Environmental Monitoring and

Assessment 63 :95-113. Hirsch, R. M., and J. R. Slack. 1984. A nonparametr ic trend test for seasonal data with serial dependence. Water Re sources Research 20 :727-732. Hirsch, R. M., J. R. Slack, and R. A. Smith. 1982. Techniques of trend analysis for monthly water quality analysis. Wa ter Resources Research 18 :107-121. Holland, D. m., P. P. Principe, and J. E. Sick les II. 1999. Trends in at mospheric sulfur and nitrogen species in the eastern United States for 1989-1995. Atmospheric Environment

33 :37-49. Husar, R. B., T. J. Sullivan, and D. F. Charles. 1991. Historical trends in atmospheric sulfur deposition and methods for assessing long-term trends in surface water chemistry. Pages

65-82 in D. F. Charles, editor. Acidic Depos ition and Aquatic Ecosystems: Regional Case Studies. Springer-Verlag, New York. Jaworski, N. A., R. W. Howarth, and L. I. Hetling. 1997. Atmospheric deposition of nitrogen oxides onto the landscape contributes to coas tal eutrophication in the northeast United States. Environmental Science and Technology Scope 31 :1995-2004. Jeffries, D. S., R. G. Semkin, F. D. Beall, and J. Franklyn. 2002. Temp oral trends in water chemistry in the Turkey Lakes waters hed, Ontario, Canada, 1982-1999. Water Air and

Soil Pollution Focus 2 :5-22. Kahl, J. S., S. A. Norton, C. A. Cronan, I. J. Fernandez, L. C. Bacon, and T. A. Haines. 1991. Maine. Pages 203-235 in D. F. Charles, editor. Acidic Deposition and Aquatic Ecosystems: Regional Case Studies . Springer-Verlag, New York, NY. Kahl, J. S., S. A. Norton, I. J. Fernandez, K. J. Nadelhoffer, C. T. Driscoll, and J. D. Aber. 1993a. Experimental inducement of nitrogen saturation at the watershed scale. Environmental Science and Technology 27 :565-568. Kahl, J. S., S. A. Norton, T. A. Haines, and R. B. Davis. 1993b. Recent trends in the acid-base status of surface waters in Maine, U.S.A. Water Air and Soil Pollution 67 :281-300. Kahl, J. S., S. A. Norton, T. A. Haines, C. W. Fay, and R. B. Davis. 1989. The influence of organic acidity on the chemistry of Maine surface waters. Water Air and Soil Pollution

46 :221-233.74 Kahl, J. S., S. A. Norton, T. A. Haines, E. A. Rochette, R. H. Heath, and S. C. Nodvin. 1992. Mechanisms of episodic acidification in low-order streams in Maine, USA.

Environmental Pollution 78 :37-44. Kahl, J. S., and M. Scott. 1994. High Elevat ion Lake Monitoring in Maine: 1986-89. Maine Department of Environmental Protection, Augusta, Maine. Kaufmann, P. R., A. T. Herlihy, J. W. Elwood, M. E. Mitch, W. S. Overton, M. J. Sale, J. J. Messer, K. A. Cougan, D. V. Peck, K. H. Reck how, A. J. Kinney, S. J. Christie, D. D. Brown, C. A. Hagley, and H. I. Jager. 1988. Ch emical Characteristics of Streams in the

Mid-Atlantic and Southeastern United States . Volume I: Population Descriptions and Physico-Chemical Relationships. U.S. Envi ronmental Protection Agency, Washington,

D.C. Keene, W. C., and J. N. Galloway. 1984. Organic acidity in precipitatio n of North America. Atmospheric Environment 18 :2491-2498. Keller, W., S. S. Dixit, and J. Heneberry. 2001. Ca lcium declines in northeastern Ontario lakes.

Canadian Journal of Fish eries and Aquatic Science 58 :2011-2020. Kingston, J. C., and H. J. B. Birks. 1990. Dissolv ed organic carbon recons tructions from diatom assemblages in PIRLA project lakes, North Am erica. Philosophical Transactions of the Royal Society of London 327 :279-288. Krug, E. C., and C. R. Frink. 1983. Acid rain on acid soil: A new perspective. Science 221 :520-

525. Landers, D. H., J. M. Eilers, D. F. Brakke, W. S. Overton, P. E. Kellar, W. E. Silverstein, R. D. Schonbrod, R. E. Crowe, R. A. Linthurst, J. M. Omernik, S. A. Teague, and E. P. Meier.

1987. Characteristics of Lakes in the Wester n United States - Volume I: Population Descriptions and Physico-Chemical Rela tionships. U.S. Environmental Protection Agency, Washington, D.C. Landers, D. H., W. S. Overton, R. A. Linthurst, and D. F. Br akke. 1988. Eastern Lake Survey: Regional estimates of lake chemistry. Environmental Science and Technology 22 :128-

135. Larsen, D. P., K. W. Thornton, N. S. Urquhart , and S. G. Paulsen. 1994. The role of sample surveys for monitoring the condi tion of the nation's lakes. Environmental Monitoring and

Assessment 32 :101-134. Larsen, D. P., and N. S. Urquhart. 1993. A framework for assessing the sensitivity of the EMAP design. Pages 4.1-4.37 in D. P. Larsen and S. J. Christie, editors. EMAP-Surface Waters 1991 Pilot Report. U.S. Environmental Protection Agency, Corvallis, OR. Lawrence, G. B. 2002. Persistent episodic acidification of streams linked to acid rain effects on soil. Atmospheric Environment 36 :1589-1598. Lawrence, G. B., M. B. David, G. M. Lovett, P. S. Murdoch, D. A. Burns, J. L. Stoddard, B. P. Baldigo, J. H. Porter, and A. W. Thompson. 1999. Soil calcium status and the response of stream chemistry to changing acidic depos ition rates. Ecological Applications 9 :1059-

1072. Likens, G. E., F. H. Bormann, and N. M. Johnson. 1972. Acid rain. Environment 14 :33-40. Likens, G. E., C. T. Driscoll, and D. C. Bus o. 1996. Long-term effects of acid rain: Response and recovery of a forest ecosystem. Science 272 :244-246. Likens, G. E., E. S. Edgerton, and J. N. Galloway. 1983. The composition and deposition of organic carbon in pr ecipitation. Tellus 35 :16-24.75 Lindberg, S. E., J. M. Coe, and W. A. Hoffma n. 1984. Dissociation of weak acids during Gran plot free acidity titrations. Tellus 36 :186-191. Linthurst, R. A., D. H. Landers, J. M. Eilers, D. F. Brakke, W. S. Overton, E. P. Meier, and R. E.

Crowe. 1986a. Characteristics of lakes in the eastern United St ates - Volume I: Population Descriptions, and Physico-Chem ical Relationships. U.S. Environmental Protection Agency, Washington, D.C. Linthurst, R. A., D. H. Landers, J. M. Eilers, D. F. Brakke, W. S. Overton, E. P. Meier, and R. E.

Crowe. 1986b. Characteristics of Lakes in th e Eastern United States. Volume I. Population Descriptions and Physico-Chem ical Relationships. U.S. Environmental Protection Agency, Washington, D.C. Loftis, J. C., and C. H. Taylor. 1989. Detecting ac id precipitation impacts on lake water quality. Environmental Management 13 :529-539. Lynch, J. A., V. Bowersox, and J. Grimm. 2000. Acid rain reduced in the eastern U.S. Environmental Science and Technology 34 :940-949. Mattson, M. D., P. J. Godfrey, M. Walk, P. A. Kerr, and O. T. Zajicek. 1997. Evidence of recovery from acidification in st reams. Water Air and Soil Pollution 96 :211-232. Miller, P. 1999. Emissions-related Acidic Deposi tion Trends in Maine. Final report for the Maine Ecological Assessment Project, EPA Of fice of Air and Radi ation, Washington,

DC. Mitchell, M. J., C. T. Driscoll, J. S. Kahl, G. E. Likens, P. S. Murdoch, and L. H. Pardo. 1996.

Climatic control of nitrate loss from forested watersheds in the nor theast United States. Environmental Science and Technology 30 :2609-2612. Murdoch, P. S., and J. L. Stoddard. 1992. The role of nitrate in the acidifi cation of streams in the

Catskill Mountain of New Yor k. Water Resources Research 28 :2707-2720. Murdoch, P. S., and J. L. Stoddard. 1993. Chemical characteristics and tem poral trends in eight streams of the Catskill Mountains, Ne w York. Water Air and Soil Pollution 67 :367-395. National Acid Precipitation Assessment Program. 1998. NAPAP Biennial Report to Congress: An Integrated Assessment. National Acid Precipitation Assessment Program, Silver Spring, Maryland. National Atmospheric Deposition Program/Nati onal Trends Network. 2002. NADP/NTN Annual Data Summary. Precipitation Chemistry in the United States. Illinois State Water Survey, Champaign, Illinois. Newell, A. D., C. F. Powers, and S. J. Ch ristie. 1987. Analysis of Data from Long-term Monitoring of Lakes. U.S. Environmen tal Protection Agency, Corvallis, OR. Nielsen, M. G., and J. S. Kahl. in press. Nutrient export to estuaries as a function of land-use and fire history. Journal of the Amer ican Water Resources Association. Norton, S. A., B. J. Cosby, I. J. Fernandez, J. S. Kahl, and M. R. Church. in press. Inter-annual

and seasonal variations in soil-air CO 2 cause variable watershed alkalinity generation. Hydrologic and Earth Science Systems. Norton, S. A., J. S. Kahl, I. J. Fernandez, L. E. Rustad, T. A. Haines, S. C. Nodvin, J. P. Scofield, T. C. Strickland, P. J. Wigington, and J. Lee. 1999. The Bear Brook Watershed in Maine (BBWM). Environmental Mon itoring and Assessment 55 :7-51. Norton, S. A., J. S. Kahl, I. J. Fernandez, L. E. Rustad, J. P. Scofield, and T. A. Haines. 1994. Response of the West Bear Brook watershed, Maine, USA, to the addition of (NH 4 ) 2 SO 4 : 3-year results. Forest Ecology and Management 68 :61-73.76 Oden, S. 1968. The acidification of air and precip itation and its conseque nces in the natural environment. Swedish National Research Council, Stockholm. Paerl, H. W. 2002. Connecting atmospheric nitr ogen deposition to coastal eutrophication. Environmental Science and Technology August 1, 2002 :232A-326A. Palmer, S. M., and C. T. Driscoll. 2002. Declin e in mobilization of toxic aluminium. Nature

417 :242-243. Paulsen, S. G., D. P. Larsen, P. R. Kaufmann, T. R. Whittier, J. R. Baker, D. V. Peck, J. McGue, R. M. Hughes, D. McMullen, D. Stevens, J. L. Stoddard, J. Lazorchak, W. Kinney, A. R. Selle, and R. Hjort. 1991. EMAP-Surface Wa ters Monitoring and Research Strategy.

Fiscal Year 1991. U.S. Environmenta l Protection Agency, Corvallis, OR. Peck, D. V. 1992. Environmental Monitoring an d Assessment Program: Integrated Quality Assurance Project Plan for the Surface Waters Resource Group. EPA/600/X-91/080, U.S. Environmental Protection Agency. Riley, J. P., and R. Chester. 1971. Introduction to Marine Chemistry. Academic Press, New

York. Rosenqvist, I. T. 1978. Acid precipitation and othe r possible sources for acidification of rivers

and lakes. Science of the Total Environment 10 :271-272. Rustad, L. E., J. S. Kahl, S. A. Norton, and I. J. Fernandez. 1994. Underestimation of dry deposition by throughfall in mixed norther n hardwood forests. Journal of Hydrology

162 :319-336.

SAS Institute Inc. 1988. SAS/STAT User's Guid e, Release 6.03 Edition. SAS Institute, Inc., Cary, NC. Schindler, D. W. 1988. The effects of acid ra in on freshwater ecosystems. Science 239 :149-157. Schindler, D. W., K. H. Mills, D. F. Malley, D. L. Findlay, J. A. Shearer, I. J. Davies, M. A. Turner, G. A. Linsey, and D. R. Cruiks hank. 1985. Long-term ecosystem stress: The effects of years of experimental acidification on a small lake. Science 228 :1395-1401. Schofield, C. L. 1976. Acid precipit ation: Effects on fish. Ambio 5 :228-230. Schofield, C. L. 1993. Habitat suitability for Brook Trout (Salvelinus fontinalis ) reproduction in

Adirondack lakes. Water Resources Research 29 :875-879. Sen, P. K. 1968. On a class of aligned rank order tests in two-way layouts. Annals of

Mathematics and Statistics 39 :1115-1124. Shannon, J. D. 1999. Regional trends in wet depos ition of sulfate in the United States and SO 2 emissions from 1980 through 1995. Atmospheric Environment 33 :807-816. Sisterson, D. L., V. C. Bowersox, A. R. Olse n, T. P. Meyers, and R. L. Vong. 1990. Deposition monitoring: methods and results. State- of-Science/Technology Report No. 6, National Acid Precipitation Assessme nt Program, Washington, DC. Skjelkvåle, B. L., J. L. Stoddard, and T. Anders en. 2001. Trends in surface water acidification in Europe and North America (1989-19 98). Water Air and Soil Pollution 130 :787-792. Smith, R. A. 1872. Air and Rain: The Beginnings of a Chemical Climatology. Longmans, Green,

London. Smol, J. P., B. F. Cumming, A. S. Dixit, and S. S. Dixit. 1998. Tracki ng recovery pattens in acidified lakes: a paleolimnological perspective. Restoration Ecology 6 :318-326. Stoddard, J. L. 1990. Plan for Converting the NAPAP Aquatic Effects Long-Term Monitoring (LTM) Project to the Temporally Integrated Monitoring of Ecosystems (TIME) Project. Internatl Report U.S. Environmental Protection Agency, Corvallis, OR.77 Stoddard, J. L. 1994. Long-term changes in waters hed retention of nitrogen: its causes and aquatic consequences. Pages 223-284 in L. A. Baker, editor. E nvironmental Chemistry of

Lakes and Reservoirs. American Chemical Society, Washington, DC. Stoddard, J. L., C. T. Driscoll, J. S. Kahl, a nd J. Kellogg. 1998a. A regional analysis of lake

acidification trends for the northeaster n U.S., 1982-1994. Environmental Monitoring and

Assessment 51 :399-413. Stoddard, J. L., C. T. Driscoll, S. Kahl, a nd J. Kellogg. 1998b. Can site-specific trends be extrapolated to a region? An acidification example for the Northeast. Ecological Applications 8 :288-299. Stoddard, J. L., D. S. Jeffries, A. Lükewille, T. A. Clair, P. J. Dillon, C. T. Driscoll, M. Forsius, M. Johannessen, J. S. Kahl, J. H. Kellogg, A. Kemp, J. Mannio, D. Monteith, P. S. Murdoch, S. Patrick, A. Rebsdorf, B. L. Skje lkvåle, M. Stainton, T. Traaen, H. van Dam, K. E. Webster, J. Wieting, and A. Wilander. 1999. Regional trends in aquatic recovery

from acidification in North America and Europe. Nature 401 :575-578. Stoddard, J. L., and J. H. Kellogg. 1993. Trends and patterns in lake acidific ation in the state of Vermont: evidence from the Long-Term Monito ring project. Water Air and Soil Pollution

67 :301-317. Stoddard, J. L., A. D. Newell, N. S. Urquhart, and D. Kugler. 1996. The TIME project design: II. Detection of regional acidificati on trends. Water Resources Research 32 :2529-2538. Sullivan, T. J. 1990. Historical Changes in Surf ace Water Acid-Base Chemistry in Response to Acidic Deposition. NAPAP State of Scie nce and Technology Report 11, National Acid Precipitation Assessment Program. Turner, R. S., R. B. Cook, H. Van Miegroet, D. W. Johnson, J. W. Elwood, O. P. Bricker, S. E. Lindberg, and G. M. Hornberger. 1990. Waters hed and Lake Processes Affecting Surface Water Acid-Base Chemistry. State-of-Sci ence/Technology Report No. 10, National Acid Precipitation Assessment Pr ogram, Washington, DC. U.S. Environmental Protection Agency. 2002. Clean Air Status and Trends Network (CASTNet): 2000 Annual Report. U.S. Environmental Prot ection Agency, Research Triangle Park,

NC. Urquhart, N. S., S. G. Paulsen, and D. P. Larsen. 1998. Monitoring for regional and policyrelevant trends over time . Ecological Applications 8 :246-257. Webb, J. R., B. J. Cosby, J. N. Galloway, and G. M. Hornberger. 1989. Acidification of native brook trout streams in Virginia. Water Resources Research 25 :1367-1377. Webb, J. R., F. A. Deviney, J. N. Galloway, C. A. Rinehart, P. A. Thompson, and S. Wilson.

1994. The Acid-Base Status of Native Brook Trout Streams in the Mountains of Virginia. Dept. Environmental Sciences , University of Virginia. Webster, K. E., P. L. Brezonik, and B. J. Ho ldhusen. 1993. Temporal trends in low alkalinity lakes of the Upper Midwest (1983- 1989). Water Air and Soil Pollution 67 :397-414. Webster, K. E., and P. O. Brezonik. 1995. Clim ate confounds detection of chemical trends related to acid deposition in Upper Midwest lakes in the USA. Water Air and Soil

Pollution 85 :1575-1580. Webster, K. E., A. D. Newell, L. A. Baker, and P. L. Brezonik. 1990. Climatically induced rapid acidification of a softwater seepage lake. Nature 347 :374-376. Wigington Jr., P. J. 1999. Episodic acidification: causes, occurrence and significance to aquatic resources. Pages Chapter 1, pg. 1-5 in J. R. Drohan, editor. The Effects of Acidic78 Deposition on Aquatic Ecosystems in Pe nnsylvania. 1998 PA Acidic deposition Conference. Environmental Resources Research Institute, University Park, PA. Wigington, P. J., J. P. Baker, D. R. DeWalle, W. A. Kretser, P. S. Murdoch, H. A. Simonin, J. Van Sickle, M. K. McDowell, D. V. Peck, and W. R. Barchet. 1996. Episodic

acidification of small streams in the nor theastern United States : Episodic Response Project. Ecological Applications 6 :374-388. Williams, M. W., J. S. Baron, N. Caine, R. Sommerfield, and R. Sanford. 1996. Nitrogen saturation in the Rocky Mountains. E nvironmental Science and Technology 30 :640-646. Wright, R. F. 1983. Predicting Acidification of North American Lakes. Acid Rain Research Report 4/1983

. Wright, R. F., E. Lotse, and A. Semb. 1993. RAIN Project: Results after 8 years of experimentally reduced acid deposition to a whole catchment. Canadian Journal of Fisheries and Aquatic Science 50 :258-268. Young, T. C., and J. L. Stoddard. 1996. The TIME proj ect design: I. Classification of Northeast lakes using a combination of geographic, hydr ogeochemical, and multivariate techniques.

Water Resources Research 32 :2517-2528.