600R96123
Assessment of Deposition Levels of
Sulfur and Nitrogen Required to
Protect Aquatic Resources
in Selected Sensitive Regions
of North America
Final Report
TU	
Technical Resources. Inc.

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EPA ERL-Corvallis Library ^
00006857
Assessment of Deposition Levels of
Sulfur and Nitrogen Required to
Protect Aquatic Resources
in Selected Sensitive Regions
of North America
Final Report
Prepared by:
Timothy J. Sullivan and Joseph M. Eilers
E&S Environmental Chemistry, Inc.
P.O. Box 609
Corvallis, OR 97339
March 31,1994
Prepared for:
Technical Resources, Inc.
3202 Tower Oaks Blvd., Suite 200
Rockville, MD 20852
Under Contract No. 68-C0-0021 (Option III) to:
U.S. Environmental Protection Agency
200 SW 35th Street
Corvallis, OR 97333
Library
U.S. Environmental Protection Agency
National Health end Environmental
Effects Beseazeh Laboratory
200 S.W. 35th Street
Corvallis, Oregon 97333

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TABLE OF CONTENTS
EXECUTIVE SUMMARY		iv
I.	BACKGROUND AND OBJECTIVES 		1
A.	NEED FOR ACID DEPOSITION STANDARDS 		1
B.	LEGISLATIVE MANDATE 	 		5
C.	OBJECTIVES		6
II.	SELECTION OF REGIONS OF CONCERN			8
A.	CRITERIA FOR SELECTION	 8
B.	FACTORS THAT INFLUENCE SENSITIVITY TO ACIDIC DEPOSITION 		9
C.	REGIONAL DELINEATION 		12
1.	West 		13
2.	Upper Midwest		15
3.	Northern Florida		17
4.	Southeastern Canada 			17
III.	APPROACH			20
A.	ACID SPECIES/PRECURSORS	 	 		20
1.	Sulfur				21
2.	Nitrogen					22
B.	TEMPORAL RESPONSE		30
1.	Chronic Acidification		30
2.	Episodic Acidification 		33
IV.	REGIONAL ASSESSMENT 		35
A.	WEST	:		 .	35
1.	Characteristics and Sensitivity to Acid Deposition		35
2.	Current and Projected Future Deposition 		36
3.	Current Surface Water Chemistry		37
4.	Quantitative Assessment of Acidification			56
a.	Monitoring		56
b.	Paleolimnology 		57
c.	Process models			58
B.	UPPER MIDWEST			59
1.	Characteristics and Sensitivity to Acid Deposition		59
2.	Current and Projected Future Deposition 		61
3.	Current Surface Water Chemistry		62
4.	Quantitative Assessment of Acidification	 			64
a.	Space-for-time substitution 			65
b.	Monitoring	 		70
c.	Paleolimnology 			71
d.	Process models . 			73
e.	Experimental Manipulation (Little Rock Lake)		73
C.	NORTHERN FLORIDA 		74
1.	Characteristics and Sensitivity to Acid Deposition		74
2.	Current and Projected Future Deposition	 		76
3.	Current Surface Water Chemistry			77
4.	Quantitative Assessment of Acidification		79

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a.	Monitoring		80
b.	Paleolimnology 		80
c.	Process models		84
D.	EASTERN CANADA		86
1.	Characteristics and Sensitivity to Acid Deposition		86
2.	Current and Projected Future Deposition 		87
3.	Current Surface Water Chemistry		88
4.	Quantitative Assessment of Acidification		90
a.	Monitoring				90
b.	Paleolimnology 		95
c.	Experimental Manipulation 			101
E.	COMPARISONS WITH DATA FROM THE ADIRONDACK MOUNTAINS 		102
V.	ASSESSMENT OF SULFUR AND NITROGEN DEPOSITION LEVELS REQUIRED TO
PROTECT AQUATIC RESOURCES 			108
A.	REGIONAL DOSE/RESPONSE ASSESSMENT		108
1.	West '	 			108
2.	Upper Midwest	 		110
3.	Florida			113
4.	Eastern Canada 		116
B.	RECOMMENDATIONS			120
1. Quantification of Chronic Acidification Responses		120
a. Standards for Sulfur and Nitrogen		128
C.	UNCERTAINTIES	 		138
1.	Land Use			140
2.	Pollutant Interactions			141
3.	Temporal Considerations			142
VI.	SUMMARY AND CONCLUSIONS				144
VII.	LITERATURE CITED			148
APPENDIX A. Assessment Methods for Evaluation of Effects on Aquatic Resources
Due to Changes in Atmospheric Depostion 	
173

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EXECUTIVE SUMMARY
There has been a growing international recognition that air pollution effects, particularly from
sulfur and nitrogen, may in some cases necessitate emission controls to reduce atmospheric
deposition. Measures to reduce emissions must rely on known or estimated dpse/response
relationships which reflect the tolerance of natural ecosystems to various inputs of atmospheric
pollutants. This need has stimulated interest in evaluating the efficacy of establishing one or
more standards for acid deposition. The Clean Air Act Amendments of 1990 (CAAA) also
included requirements to assess the effectiveness of the mandated emissions controls via
periodic assessments, and to submit to Congress an EPA report on the feasibility of adopting one
or more acid deposition standards. The purpose of this report is to provide technical information
required for assessing the feasibility of adopting one or more acid deposition standards for the
protection of aquatic resources.
We have identified four regions for inclusion in this assessment: western and upper
midwestern United States, Florida, and southeastern Canada. For the regions identified as having
sensitive or critically sensitive aquatic resources, relevant information has been compiled and
evaluated regarding the relationship between deposition loading (N and S) and the estimated (or
expected) extent, magnitude, and timing of aquatic effects. The general approach we have
employed for this task involved a "weight of evidence" evaluation of the relationships between
deposition and effects, as followed by NAPAP in the Integrated Assessment.
Sulfate is the most important anion, on a quantitative basis, in acidic deposition in most parts
of the United States. Consequently, sulfate and the controls on its inputs and processing have
-received the greatest scientific and policy attention to date, and the response of watersheds to
sulfur inputs, particularly chronic effects on surface water quality, are now reasonably well
understood.
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Nitrate in snowmelt runoff has been recognized for some time as an important component of
biological damage resulting from atmospheric deposition. Nitrate is the principal acid anion in
snowmelt in many areas of the northeastern and western United States. A number of factors can
be involved in controlling the loss of nitrogen from a forested watershed, however, including
atmospheric inputs, forest stand age and condition, soil nitrogen pools, and flowpaths of
percolation and melt-water within the catchment. Except in cases of excessive nitrogen
saturation, the effects of nitrogen deposition on surface waters are expected to be primarily
episodic in nature. Unfortunately, data required to make regional assessments of episodic effects
are generally not available. Such data need to be collected on an intensive schedule and must
include sample periods during late winter and early spring when snowmelt often causes the most
severe nitrogen-driven episodes of surface water acidification.
The acid-base chemistry of surface waters typically exhibits substantial intra- and interannual
variability. Seasonal variability in the concentration of key chemical parameters often varies by
more than the amount of acidification that might occur in response to acidic deposition. Such
variability makes quantification of acidification and recovery responses difficult, and also
complicates attempts to evaluate sensitivity to acidification based solely on "index" chemistry, as
is typically collected in synoptic lake or stream surveys. Seasonal variability is particularly
problematic in the assessments of standards for nitrogen.
Episodic acidification is nearly ubiquitous in drainage waters. Lakes and streams that have
been studied throughout the United States, Canada, and Europe nearly all experience loss of
ANC during hydrologic events. Chemical changes during episodes are controlled by a number of
natural processes, including dilution of base cation concentrations, nitrification, flushing of organic
acids from terrestrial to aquatic systems, and the neutral salt effect. Acidic deposition can also
contribute to episodic acidification, particularly via enhanced N03' leaching. Under some
conditions, episodes can also be partially caused by increased S042' concentration. There is also

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the possibility that chronic acidification by acid deposition can pre-condition a watershed, thereby
increasing the severity of episodic acidification.
Western United States
The areas, containing low-ANC lakes in the West are confined primarily to the higher elevation
mountainous regions, most of which have been glaciated. We sought to consolidate the study
areas in the West into major lake populations located within similar geomorphic units, and have
focused on the following subpopulations:
•	Sierra Nevada, CA
•	Cascades, CA, OR, WA
•	Idaho Batholith, ID, MT
•	Northwest Wyoming
•	Colorado Rockies
Portions of the mountainous West are similar to the highly sensitive areas in Norway where
watersheds contain large areas of exposed bedrock, with little soil or vegetative cover to
neutralize acidic inputs. This is particularly true of alpine regions of the Sierra Nevada, northern
Washington Cascades, the Idaho batholith, and portions of the Rocky Mountains in Wyoming and
Colorado. The current chemistry of surface waters in the West is based almost exclusively on
synoptic data from the Western Lake Survey and a small number of more localized studies.
Comprehensive assessments of lake chemistry in the West indicate that there are many low-ANC
systems, but virtually no chronically acidic waters.
Because of the proximity of well-defined population centers and industrial pollution sources in
the West to sensitive aquatic receptors in individual mountain ranges, it is important to evaluate
changes in emissions in the immediate vicinity and up-wind of sensitive resources, in addition to
region-wide assessment of emissions and deposition trends. Such data are generally not
available.

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Nitrate concentrations are virtually undetectable in the majority of western lakes in the fall.
However, in some cases, fall nitrate concentrations were surprisingly high. For example, nearly
one fourth of the lakes in NW Wyoming had N03 > 5 fieq L"1 and almost 10% had NOa' > 10
,ueq L'\ The Sierra Nevada and Colorado Rockies subregions also exhibited many lakes with
higher NO,' concentrations than would be expected for fall samples. In both areas about 10% of
the lakes had N03 concentrations above 5 ^eq L'\
Episodic acidification is clearly an important issue for surface waters throughout high-
elevation areas of the West. A number of factors pre-dispose western systems to potential
episodic effects, including:
1.	the abundance of dilute to ultradilute lakes, exhibiting very low concentrations of base
cations, and therefore ANC, throughout the year,
2.	large snowpack accumulations at the high elevation sites, thus causing substantial
episodic acidification via the natural process of base cation dilution, and
3.	short retention times for many of the high-elevation drainage lakes, thus enabling
snowmelt to rapidly flush lake basins with highly dilute meltwater.
The Sierra Nevada are particularly sensitive to potential acidic deposition effects because of
the predominance of granitic bedrock, thin acidic soils, large amounts of precipitation, coniferous
vegetation, and extremely dilute waters. Similarly, Cascade and Rocky Mountain lakes are highly
sensitive to potential acidic deposition effects. There are no data to suggest that lakes in these
areas have experienced chronic acidification to ANC values less than zero to date, and based on
examination of current chemistry it appears that chronic acidification has not occurred to any
significant degree. It is possible, however, that episodic effects have occurred under current
deposition regimes, and that N03° concentrations have caused a small loss of ANC on a chronic
basis at many high-elevation sites. Unfortunately, the data that would be needed for such
determinations have not been collected to a sufficient degree in acid-sensitive areas of the West
to permit any regional assessment of either episodic or chronic N-driven acidification.

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It is our best professional judgement that a reasonable standard for protecting sensitive lakes
throughout large areas of the West from adverse effects of chronic sulfur deposition is near
10 kg S ha'1 yr'1. In some mountainous areas of the West, however, where highly dilute lakes are
numerous, such a standard would be considerably lower, likely in the range of 2 to 5
kg S ha1 yr1.
Use of a deposition standard in the West may be problematic in certain areas because of the
wide range in precipitation volume. In some areas of the Pacific Northwest, precipitation volumes
are so great (up to 4 m) that all but the most conservative standards will be violated under natural
conditions. Some consideration needs to be given to an annual-weighted pollutant concentration
standard or a combination deposition/concentration standard rather than a stand-alone deposition
standard in many areas of the west.
Upper-Midwestern United States
The Upper Midwest is characterized by numerous lakes created by repeated glaciations. The
region shows little topographic relief and the deep glacial overburden results in little or no
exposed bedrock. Sensitive aquatic resources in the Upper Midwest are largely seepage lakes.
Those seepage lakes with low base cation concentrations receive nearly all of their hydrologic
inputs as precipitation directly on the lake surface. Consequently, these lakes generally have
long hydraulic residence times, thus providing an opportunity for in-lake reduction and
assimilation processes to neutralize much of the acidic inputs which would otherwise be
concentrated from evaporation.
Emissions of S02 and NOx in the Upper Midwest appear to have increased dramatically in the
20th century. Recent trends indicate that emissions reached a maximum in about 1978 and have
decreased since that period. The Upper Midwest has a large population of low ANC lakes, but
relatively few acidic (ANC < 0) lakes. Paleolimnological evidence suggests slight

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acidification of selected lakes, consistent with the modest historical and current rates of sulfur
deposition.
Concentrations of inorganic nitrogen are uniformly low throughout the Upper Midwest and
are efficiently retained in both terrestrial and aquatic systems. Snowmelt does not provide any
significant nitrate influx to lakes in the Upper Midwest because much of the snowmelt first
infiltrates into the soil prior to reaching the lakes. The key issue in this region, therefore, appears
to be chronic, sulfur-driven acidification.
The Upper Midwest is a diverse region, with pronounced spatial gradients from west to east
in hydrologic lake type, major ion chemistry, and sulfur deposition. Lakes within the easternmost
portion of the region receive the highest sulfur deposition and also exhibit the greatest sensitivity
to potential acidification. Although deposition in the Upper Midwest has been declining in recent
years, those lakes with ANC near zero could be expected to acidify if trends in S emissions
reversed.
Although paleolimnological data suggest that some upper midwestern lakes have acidified
since pre-industrial time, there is little paleolimnological evidence indicating substantial
widespread acidification in this region. It is likely that land use changes and other human
perturbations of upper midwestern lakes and their watersheds have exerted more influence on the
acid-base chemistry of lakes than has acidic deposition. This result is not unexpected because
acidic deposition has been much smaller in magnitude in the Upper Midwest than in most areas
of the eastern United States. It is clear, however, that the portion of the region most likely to have
experienced acidification from acidic deposition is the Upper Peninsula of Michigan, where acidic
seepage lakes are particularly numerous; acidic deposition is highest for the region, and the
lakewater [S042")/[CB] ratio is commonly > 1.0. The percentage of acidic lakes in the eastern
portion of the Upper Midwest region is 18% to 19%. In our judgement, a reasonable sulfur
standard for the most sensitive aquatic resources in the Upper Midwest can be approximated by
current levels of deposition in the eastern portion of the region, about 5 kg S ha"1 yr"\ In the

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western parts of this region, surface waters are less sensitive to sulfur deposition effects, and an
appropriate standard would be much higher. Collection of additional paleolimnological data in
this area would enhance our ability to draw quantitative conclusions for the region as a whole.
Florida
Florida lakes are located in marine sands overlying carbonate bedrock. The origin of many
Florida lakes is from dissolution of the limestone and subsequent collapse of the overlying soils
into the cavity (sinkhole formation). Where groundwater interactions with the deeper aquifers are
present, surface waters can be highly alkaline. However, those lakes with hydrologic
contributions from shallow aquifers in highly weathered sands can be quite acidic and
presumably sensitive to acidic deposition. As is the case elsewhere, the key to understanding the.
potential response of Florida lakes to acid inputs is related largely to knowledge of the hydrologic
flow paths. Northern Florida contains one of the largest populations of acidic lakes in the United
States. Seventy-five percent of the Panhandle lakes are acidic, as are 26% of the lakes in the
northern peninsula.
Evidence for acidification of some Florida lakes has been supported by historical analyses of
lake chemistry, inferred historical deposition, and paleolimnological reconstructions of lake pH.
However, the case for acidification by acid deposition is equivocal and the interpretation is
complicated by profound regional and local changes in land use and hydrology. An alternative
explanation (other than acidic deposition) for the apparent acidification of some lakes in Florida is
the regional decline in the potentiometric surface of the groundwater. Large groundwater
withdrawals of the Floridan aquifer for residential and agricultural purposes may have contributed
to reduced groundwater inflow of base cations into seepage lakes, thereby causing lakewater
acidification. Superimposed on the complex heterogeneity of Florida lakes is a high incidence of
anthropogenic disturbance, mostly related to agriculture. Besides the increased atmospheric

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deposition in Florida in the 1950s, other changes have also occurred. The human population has
increased markedly.
Historical changes in Florida lakewater chemistry, as inferred from diatoms, showed a distinct
geographical pattern. All five of the paleolimnological study lakes in the Trail Ridge region
showed some evidence of acidification, some strongly linked in timing to both the period of
increasing acidic deposition and increased withdrawals from regional aquifers. Trail Ridge lakes
showed diatom-inferred ApH ranging from -0.20 to -0.91. No clear evidence of acidification was
observed for lakes in Ocala National Forest (three lakes) or the Panhandle (eight lakes), except
Lake Five-O, where gross hydrological change was implicated. It is most likely that several
factors have caused the recent acidification of lakes in the Trail Ridge area suggested by the
diatom data. Acidic deposition is implicated, but changing lake stage and the linked
phenomenon of evapoconcentration also may be important. Thus, it is not clear whether current
levels of sulfur deposition have caused recent acidification of lakes in Florida. If such acidification
has occurred, it has likely been restricted to a relatively small geographic area, in the Trail Ridge
region.
Eastern Canada
A large percentage of Canadian paleolimnological studies have focused on lakes near large
point sources, near mining effluents, or on lakes that have been manipulated (limed or acidified).
Thus, a regional paleolimnological data base with which to evaluate acidification associated with
long-range transported pollutants is not available in Canada. In regions where acidic precipitation
occurs, however, recent pH declines have been inferred for at least some lakes.
Data collected from lakes in the vicinity of the smelting operations near Sudbury, Ontario,
have provided a wealth of information with which to quantify both acidification and recovery
responses. Particularly useful data have included the results from both paleolimnological and
monitoring/survey studies. In addition, research in the Sudbury area has provided some of the

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best data available with which to evaluate the reliability of diatom inferences of pH change.
However, interpretation of dose-response data from Sudbury is complicated by the very high
levels of acidic deposition received by many of the watersheds that have been studied. The
observed changes in deposition and associated effects, have been substantially greater than
would be experienced during any realistic future deposition scenarios for the regions under
investigation for this report.
Beyond the general vicinity of Sudbury, quantitative data on acidification response are scarce
for eastern Canada. Although current chemistry has been investigated to a considerable degree
through lake surveys, the resulting data are insufficient for quantitative dose-response
assessment. We recommend using model-based estimates of appropriate sulfur standards being
generated by ERL-C (M.R. Church, personal communication) for the Adirondack region, as a
reasonable first approximation for southeastern Canada.
Feasibility of Acid Deposition Standards
Quantification of sulfur and nitrogen dose-response relationships is difficult for the regions
considered in this report (Western and Upper Midwestern United States, Florida, and Eastern
Canada). Limited data availability precludes rigorous quantitative assessment in most cases. In
particular, data are scarce in the following categories:
•	episodic acidification, especially in the West
•	groundwater inflow (and associated neutralization) to seepage lakes
•	seasonal surface water chemistry data, particularly for nitrogen and aluminum
•	model input parameters (especially soils characteristics) for drainage systems
•	deposition (wet and dry) data at high elevation sites
•	regional paleolimnological data, especially in upper Michigan and portions of eastern
Canada

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The data that are available tor the regions under investigation for his report were, however,
compared and contrasted with more intensive data generated in other regions that have been
impacted by acidic deposition and studied in greater detail. Such comparisons were useful to
place bounds on the magnitude of the acidification response. Quantitative dose-response
relationships for sulfur have been determined, using a variety of approaches, in a number of
regions in North America and Europe. Such studies have included, for example, measured
changes in water chemistry during periods when sulfur deposition changed appreciably, regional
paleolimnological investigations, whole-catchment manipulation studies, and intensive process
modeling.
Measured changes in surface water chemistry in areas that have experienced short-term
(< 20 yr) changes in chemical constituents in response to changes in mineral acid inputs are
available from a number of sources. These include results from manipulation experiments and
changes in acidic deposition. Proportional changes in ANC (expressed as [HC03" - H+]), [CB],
and [Al|], relative to changes in [SO/] or [SO/ + N03 ], were summarized for lakes and streams
in which such changes have been measured. They include lakes in the Sudbury region of
Ontario, the Galloway lakes area of Scotland, a stream site at Hubbard Brook, New Hampshire,
and catchment manipulation experiments in the RAIN project in Norway and Little Rock Lake in
Wisconsin. Most of the observed changes are coincident with decreased acidic deposition, and it
is unclear whether acidification and recovery are symmetrical. F-factors (A [CB] A [S042' +
NOa']) in the range of 0.5 to 0.9 are apparently typical for lakes having low [CB], although lower
values (0.35 to 0.39) were observed for the highly sensitive catchments at Sogndal, Norway,
which are characterized by thin soils and much exposed bedrock, as is common in many areas of
southern Norway and the western United States. The proportional change in ANC relative to
[SO,2' + N03 ] change was variable, within the range of 0.1 to 0.5. The proportional change in Al
was smaller, ranging up to 0.15.

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Consideration of the efficacy of adopting one or more acid deposition standards for the
protection of surface water quality from potential adverse effects of sulfur and nitrogen deposition
is a multifaceted problem. It requires that sulfur and nitrogen be treated separately as potentially-
acidifying agents, and that separate estimates for each be generated for all individual, well-
defined regions or subregions of interest. Appropriate criteria must be selected as being
indicative of damaged water quality, for example ANC or pH. Once a criterion has been selected,
a critical value must be estimated, below which the criterion should not be permitted to fall. ANC
criteria have been set at 0, 20, or 50 fieq l_ 1 in various European applications. Selection of critical
values for ANC or pH is confounded by the existence of lakes and streams that are acidic or very
low in pH or ANC due entirely to natural factors, irrespective of acidic deposition. In particular,
low contributions of base cations in solution, due to low weathering rates and/or minimal contact
between drainage waters and mineral soils, and high concentrations of organic acids contribute
to naturally low pH and ANC in surface waters.
Acid deposition standards might be selected on the basis of protecting aquatic systems from
chronic acidification; conversely episodic acidification might also be considered, and would be of
obvious importance in regions where hydrology is dominated by spring snowmelt. Thus,
selection of appropriate acid deposition standards involves consideration of a matrix of factors.
Sulfur deposition is a potential concern in all of the regions under investigation for this report.
Some degree of chronic acidification attributable to sulfur deposition has occurred in southeastern
Canada, in the eastern portion of the Upper Midwest region, and possibly in the Trail Ridge
region of north-central Florida. Regional quantification of the amount of acidification that has
occurred in southeastern Canada and the Upper Midwest is not possible with existing data. In
the absence of additional regional data, we recommend that first-approximations of acid-sensitivity
for eastern Canada should be based on more quantitative results that have been obtained in the
Adirondack Mountains. It is likely that an appropriate sulfur deposition standard for eastern
portions of the Upper Midwest would be somewhat less than peak deposition values recorded in

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the 1970's, although it is not possible to quantify exactly how much less, based on available data.
Furthermore, sulfur deposition in this region has been declining steadily in recent years, and will
therefore likely be of less concern in the future than in many other regions of the country. Our
preliminary recommendation, based largely on professional judgement, is to use a sulfur standard
of 5 kg S ha'1 yr' for the most sensitive (eastern) portion of the Upper Midwest.
Based on analysis of available sulfur dose-response data for sensitive watersheds worldwide,
it is clear that proportional changes in ANC and base cations in drainage waters in response to
changes in sulfur inputs are highly variable. Documented F-factors are generally above 0.5,
although lower values have been found. Perhaps the best available estimate of an appropriate F-
factor for highly sensitive watersheds, such as are found in the western United States, would be
based on the experimental values obtained at Sogndal, in western Norway (F ~ 0.4). This alpine
watershed exhibits substantial areas of exposed bedrock, and contains shallow acidic soils. As
such, it appears to be a reasonable surrogate for highly sensitive watersheds in the West.
Assuming F = 0.4, we calculated that relatively minor increases in lakewater S042'
concentration would lead to chronic acidity (ANC < 0) in many lakes in the Sierra Nevada and
Cascade Mountain ranges. An estimated five percent of the lakes in these subregions would
become acidic with increased S042" concentration of only 27 to 30 ueq L"'. An approximate four-
fold increase in sulfur deposition in these regions, to levels in the range of 2 to 8 kg S ha' yr',
would be required to achieve such increases in lakewater S042' concentrations. In other
subregions of the West, the required S042' increase estimated to cause 5% of the lakes to
become acidic is somewhat higher (55 to 70 fieq L '), but still low compared to S042'
concentrations currently found throughout the eastern United States. Sulfur deposition levels of 5
to 10 kg S ha"' yr"' would likely cause lakewater S042' concentrations to increase to these levels.
If we base this analysis on the lowest percentile lake in the subregional ANC distribution,
increased S042' concentrations of 35 to 63 peq L1 would cause chronic acidity in the Idaho
Batholith, Wyoming, and Colorado subregions, assuming F=0.4.

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These estimates of increased S042' concentration required to acidify western lakes within the
lower percentiles of acid-sensitivity are based on fall chemistry and chronic acidification
processes. It is likely, however, that sensitive watersheds in the western United States would
experience episodic acidification (especially during snowmelt) at sulfur deposition levels lower
than those that would cause chronic acidification. In most cases, episodic pH and ANC
depressions during snowmelt are driven by natural processes (mainly base cation dilution) and
nitrate enrichment. It appears likely that sulfur deposition will also contribute to episodic
acidification of sensitive western surface waters at deposition levels below those that would cause
chronic acidification. Episodes have been so little studied within the region, however, that it is not
possible to provide quantitative estimates of episodic sulfur standards for the western subregions
of concern.
Lakewater concentrations of N03' were surprisingly high' in many high-elevation sites included
in the Western Lake Survey, despite the possible bias caused by the failure of EPA's Western
Lake Survey to collect samples at many of the highest elevation areas in the Rocky Mountains
due to frozen lake conditions. Based on existing data, it appears likely that many high-elevation
lakes in the West are currently experiencing N deposition sufficiently high to cause chronic N03"
leaching, and likely associated chronic acidification. Furthermore, it is also likely that many of
these sites that exhibit fall concentrations of N03 in the range of 10 to 30 fieq L"' have
substantially higher N03 concentrations during spring. Thus, the weight of evidence suggests
that episodic acidification associated with nitrogen deposition may be occurring to a significant
degree in many high-elevation western lakes. Unfortunately, sufficient data are not available with
which to adequately evaluate this potentially important issue.
Just as the key issues and sensitivities vary from region to region, the principal uncertainties
and weaknesses in existing data also vary across the regions selected for study. The most
significant data deficiencies and unanswered questions we have identified are as follows:

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West -
seasonal water chemistry data, especially during spring
deposition data at high elevation sites
biogeochemical data on nitrogen cycling
more detailed analyses of dose/response in individual mountain ranges
Eastern Canada - regional paleolimnological data
regional modeling (i.e., MAGIC)
statistical sampling of chronic lake and stream chemistry
seasonal water chemistry, especially during spring
Upper Midwest - regional paleolimnological data in eastern portion of the region
hydrological and geochemical studies of seepage lakes
Florida -
variable influence of acidic deposition versus groundwater withdrawal in
regulating acid-base status
hydrological and geochemical studies of seepage lakes
In addition to the primary uncertainties outlined above, and discussed in the body of this
report, there are a number of other issues that complicate determination of dose-response
relationships for sulfur and nitrogen. Acidification or recovery of surface water acid-base status in
response to changes in atmospheric deposition can be exacerbated or offset by changes in other
atmospheric factors or land use. Climatic changes (both long-term and inter-annual variability),
forest growth and management, and nitrogen saturation may have relatively large effects that can,
in some cases, overshadow catchment responses to changes in sulfur or nitrogen deposition.
Our ability to incorporate such changes into predictive models is quite limited at present, and very
little research is being conducted in the United States with which to improve the situation.

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I. BACKGROUND AND OBJECTIVES
A. NEED FOR ACID DEPOSITION STANDARDS
There has been a growing international recognition that air pollution effects, particularly from
sulfur and nitrogen, may, in some cases, necessitate emission controls to reduce atmospheric
deposition. Measures to reduce emissions must rely on known or estimated dose/response
relationships which reflect the tolerance of natural ecosystems to various inputs of atmospheric
pollutants. This need has given rise to the concepts of critical levels of pollutants and critical
loads of deposition (e.g. Bull 1992), as well as interest in evaluating the efficacy of establishing
one or more standards for acid deposition. A critical load can be defined as "a quantitative
estimate of an exposure to one or more pollutants below which significant harmful effects on
specified sensitive elements of the environment do not occur according to present knowledge"
(e.g. Nilsson 1986, Gundersen 1992). Such an approach to establishing a standard is intuitively
satisfying. However, the assignment of a standard or critical load of S or N for any particular
region may be difficult to defend scientifically. A variety of natural processes and anthropogenic
activities affect the acid-base chemistry of lakes and streams, in addition to atmospheric
deposition of S and N. Natural variability often exceeds anthropogenic acidification and
quantification of dose-response relationships is very difficult. In general, the biogeochemical
cycling of sulfur and its role in watershed acidification is better understood than is the case for
nitrogen. The N cycle itself is extremely complex and controlled by many factors besides
emission/deposition. Also, nitrogen inputs that may be beneficial to some systems may be
harmful to others. The loadings of nitrogen or sulfur that may be required to protect the most
sensitive elements of an ecosystem may be unrealistically low in terms of economic or other
considerations, and may be difficult or impossible to quantify.
The basic concept of critical load is relatively simple, as the threshold concentration of
pollutants at which harmful effects on sensitive receptors begin to occur. Implementation of the
concept is, however, not at all simple or straight-forward. Practical definitions for particular

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receptors (soils, fresh waters, forests) have not been agreed to easily. Different research groups
(nations) have employed different definitions and different levels of complexity (Bull 1991, 1992).
Constraints on the availability of suitable, high-quality, regional data have been considerable.
A number of documents have been prepared in conjunction with the United Nations
Economic Commission for Europe (UN/ECE) critical loads research efforts over the past several
years. These have included documentation of methodologies (e.g., ECE 1990) and presentation
of critical loads maps for portions of Europe. In addition, a number of other background
documents have recently been prepared in conjunction with the on-going critical loads research
efforts in Europe (e.g., Gundersen 1992, Bobbink et al. in press, Kamari et al. in press, Hessen et
al. 1992, Lovblad and Erisman 1992).
A simplistic and generalized attempt to quantify critical loads for sulfur and nitrogen was
presented at the Skokloster workshop (Nilsson and Grennfelt 1988), based on a long-term mass-
balance approach. A stable base cation pool was used as the criterion for defining the critical
load. This implied an absence of soil acidification, and allowed a connection between the critical
loads of S and N. Leaching of both N03' and S042' above the production rate of base cations via
weathering will eventually lead to soil acidification. The permissible input of nitrogen for
designation of the critical load was the amount allocated to forest growth, forest floor
accumulation, and an "acceptable" leaching of 1-2 kg N ha"1 yr"1. On this basis, Nilsson and
Grennfelt (1988) estimated critical loads of N for Europe to be in the range of 3 to 20
kg N ha"1 yr'1, depending on forest productivity.
Although some ECE working groups have been developing fairly complex, process-based
approaches, the severe constraints on data availability generally necessitate creating maps based
on the more simplistic steady-state approaches, which tend to have more substantial problems.
For example, a calculation frequently employed for estimation of the critical load of sulfur to
surface waters is based on assumed pre-industrial and current base cation fluxes (Henriksen et al.
1990a,b; 1992). There is significant uncertainty in the estimates of current base cation input,

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especially on a regional basis. It is even more difficult to quantify pre-industrial base cation
deposition.
Terminology in this research area can cause some confusion. It has been assumed that it
will not be possible to reduce loads below critical values for some sensitive systems in Europe,
and also that dynamic watershed processes cause lag periods in the acidification and recovery
responses. These problems have given rise to the concept of target loads" (e.g. Henriksen and
Brakke 1988), which implies policy relevance, rather than strictly ecological justification. "Critical
loads" and "target loads" are conceptually different. A critical load is a characteristic of a specific
environment, which can be estimated by a variety of mechanistic and empirical approaches. A
target load can be based on political, economic, or temporal considerations, and implies that the
environment will be protected to a specified level (i.e. certain degree of allowable damage) and/or
over a specified period of time. For example, a given target load may be sufficiently low as to
protect a particular ecosystem from significant environmental degradation over a ten-year period,
but in fact be substantially higher than the long-term critical load for that ecosystem. There has
been a rapid acceptance of the concept of critical and target loads throughout Europe for use in
political negotiations concerning air pollution and development of abatement strategies to mitigate
environmental damage.
Criteria of unacceptable change are typically set in relation to known effects on aquatic and
terrestrial organisms. For protection of aquatic organisms, the ANC of runoff water is most
commonly used (Nilsson and Grennfelt 1988, Henriksen and Brakke 1988, Sverdrup et al. 1990).
Critical loads of ANC. i.e., concentrations below which ANC should not be permitted to fall, have
been set at 0, 20, and 50 ,ueq L"1 for various applications (e.g., Kamari et al. 1992). Designation
of an ANC limit is confounded by natural acidification processes, which can also reduce ANC to
low, or even negative, values.
Critical loads are often determined separately for soils and surface waters, and the resulting
estimates may differ. In general, surface waters appear to be more sensitive (i.e. have lower

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critical loads) than soils within a given area (Jenkins et al. in press). Because the objective of
implementing the critical load concept is to protect the entire ecosystem from degradation, the
overall critical load for the ecosystem is the lowest critical load observed for the various sensitive
receptors. In other words, if the surface waters are protected, the soils will generally also be
protected. Critical loads will also differ from site to site depending on the inherent sensitivity of
the environment.
The mapping of critical loads throughout Europe was initiated at several international
workshops within the United Nations Economic Commission for Europe (UN/ECE). The resulting
maps assigned critical load values to discrete geographical areas (grids), and provided the basis
for comparison with current or projected atmospheric deposition. A great deal of effort has gone
into mapping activities on national and international scales in Europe during the past four years.
Such maps are intended for use in developing future pollution abatement strategies.
Nitrogen saturation and nitrate leaching have been proposed as indicators of ecosystem
stability. The definition of N saturation, and interpretation of nitrogen effects on ecosystem
stability, require the evaluation of N03 leaching data within the context of data from unaffected
areas. This is difficult in Europe because N deposition is elevated throughout most forested
regions (Gundersen 1992). Based on available data, background nitrate leaching from coniferous
forests has been estimated to be in the range of 1 to 3 kg N ha'' yr"' (e.g., Nilsson and Grennfelt
1988, Hauhs et al. 1989). Estimates in this range are currently being used in critical loads
calculations.
As a forest ecosystem approaches the point of nitrogen saturation, N03 leaching will first
become pronounced during the dormant season when vegetative uptake is low. The biological
control on N03 leaching results in a distinct seasonality in the patterns of N03 leaching from soils
and the resulting N03" concentrations in drainage waters. This biological control of N03 leaching,
and consequent seasonality in N03" output fluxes, can be eliminated as the ecosystem becomes
N-saturated. This was emphasized by Hauhs et al. (1989) who showed a progressive reduction of

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the N03" seasonality at the Lange Bramke and Dicke Bramke sites in Germany. This loss of
biological control appears to be a critical factor indicating N-saturation.
Critical loads can be evaluated on an empirical basis, using input/output budgets. For
example, Grennfelt and Hultberg (1986) examined N03 leaching across a gradient of atmospheric
N input in Europe, and found increased N03' leaching at a threshold of wet deposition input of
about 10-15 kg N ha1 yr'1. Such an empirical approach has limitations, however, because other
factors besides atmospheric input can regulate the extent of N03' leaching (Skeffington and
Wilson 1988, Gundersen 1992). Forest decline, in particular, can confound the analysis.
Gundersen (1992) emphasized that such empirical analyses can yield useful information, but
cautioned that the data should be separated by scale (plot or catchment) and ecosystem type
(coniferous or deciduous), and sites with obvious forest decline or N-fixation should be excluded.
B. LEGISLATIVE MANDATE
In 1990 the Clean Air Act was amended by Congress, in part in an effort to reduce the
perceived adverse environmental impacts of acidic deposition. Title IV of the Clean Air Act
Amendments of 1990 (CAAA) required a ten million ton reduction in annual atmospheric
emissions of sulfur dioxide and approximately a two million ton reduction in annual nitrogen oxide
emissions. The CAAA also included requirements to assess the effectiveness of the mandated
emissions controls via periodic assessments. In addition, the U.S. Environmental Protection
Agency is required by Section 404 of the CAAA to submit to Congress a report on the feasibility
of adopting one or more acid deposition standards:
"Not later than 36 months after the date of enactment of this Act,
the Administrator of the Environmental Protection Agency shall
transmit to the Committee on Environment and Public Works of the
Senate and the Committee on Energy and Commerce of the House
of Representatives a report on the feasibility and effectiveness of an
acid deposition standard or standards to protect sensitive and
critically sensitive aquatic and terrestrial resources. The study
required by this section shall include, but not be limited to,
consideration of the following matters:

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(1)	identification of the sensitive and critically sensitive aquatic and terrestrial
resources in the United States and Canada which may be affected by the
deposition of acidic compounds;
(2)	description of the nature and numerical value of a deposition standard or
standards that would be sufficient to protect such resources;
(3)	description of the use of such standard or standards in other Nations or by
any of the several States in acid deposition control programs;
(4)	description of the measures that would need to be taken to integrate such
standard or standards with the control program required by title IV of the
Clean Air Act;
(5)	description of the state of knowledge with respect to source-receptor
relationships necessary to develop a control program on such standard or
standards and the additional research that is ongoing or would be needed
to make such a control program feasible; and
(6)	description of the impediments to implementation of such control program
and the cost-effectiveness of deposition standards compared to other
control strategies including ambient air quality standards, new source
performance standards and the requirements of title IV of the Clean Air
C. OBJECTIVES
The purpose of this report is to provide technical information required for assessing the
feasibility of adopting one or more acid deposition standards for the protection of aquatic
resources. The EPA's Environmental Research Laboratory-Corvallis (ERL-C) is responsible for
components of the required report that deal with aquatic effects from acidic deposition. In
particular, ERL-C staff are focussing efforts on the identification of sensitive aquatic resources and
the quantification of relationships between deposition loadings and aquatic affects. Quantitative
model-based analyses are on-going at ERL-C for areas of the United States intensively studied in
EPA's model forecasting program, the Direct Delayed Response Project (DDRP, Church et al.
1989). The Model of Acidification of Groundwater in Catchments (MAGIC, Cosby et al. 1985a,b)
is being used to project changes in surface water chemistry for a range of sulfur and nitrogen
deposition scenarios, assuming a range of nitrogen retention efficiencies. Results of these
MAGIC model simulations will form the basis for ERL-C's assessment of the feasibility of adopting
Act.'

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one or more acid deposition standards for the regions under study: Adirondack Mountains. Mid-
Appalachians, and Southern Blue Ridge.
The report provided herein is intended to-complement ERL-C's efforts by addressing
questions related to the feasibility of an acid deposition standard or standards in other regions not
included within ERL-C's modeling efforts. Specific objectives are to:
1.	identify areas of the U.S. and Canada, other than those being modeled by ERL-C,
with sensitive aquatic resources;
2.	summarize and evaluate available information for each of the areas identified (in
#1 above) with respect to the relationship between deposition loading (N and S)
and the extent, magnitude, and timing of aquatic effects; and
3.	summarize and evaluate deposition levels sufficient to protect sensitive and
critically sensitive aquatic resources.

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II. SELECTION OF REGIONS OF CONCERN
A. CRITERIA FOR SELECTION
The areas of North America with low ANC surface waters are discontinuous, clustered into a
discrete number of regions or physiographic provinces that contain the vast majority of low ANC
lakes and streams. It was recognized relatively early that most of the major concentrations of low
ANC surface waters were probably located in areas underlain by low-weathering bedrock.
Subsequent compilations of available water chemistry data (e.g., Omernik and Powers 1982,
Omernik 1990, as shown in Eilers and Selle 1991) refined and expanded this image of sensitive
areas in North America. The extensive research program conducted through NAPAP provided
additional insight into factors contributing to our understanding of sensitive regions by revealing
the importance of soil composition and hydrologic flowpath, in addition to geology, in delineating
sensitive regions.
Broad areas in North America that contain large populations of low ANC lakes and streams
include much of southeastern Canada, portions of the northeastern United States (particularly the
Adirondack Mountains), the mid-Appalachian Mountains, northern Florida, the Upper Midwest,
and the western United States (Figure II.A.1). The Adirondack Mountains and the streams in the
mid-Appalachian Mountains include many acidified surface waters. These areas are being
evaluated in a separate report by the U.S. Environmental Protection Agency (The Nitrogen
Bounding Study, M.R. Church, personal communication). Other regions in North America for
which there are large populations of low ANC surface waters are summarized in Table II.A.1.
However, the ability to make quantitative appraisals of the feasibility of defining acidic deposition
standards is limited in some of these areas because of the relatively restricted scope of studies
that have been conducted in these regions and the paucity of related data on factors that
contribute to watershed sensitivity (e.g. soils data, hydrology). Furthermore, some of these areas
lack sufficient information on deposition chemistry, and some of the sensitive aquatic resources
are located in such remote areas, lacking sufficient up-wind pollution sources, that the need for

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Figure II.A.1. Major areas of North America containing low-ANC surface waters. Not shown are
portions of western Canada and Alaska. Sensitive areas are depicted by shading.
(Source: Charles 1991)
atmospheric deposition standards in the near future seems unlikely. In view of these issues, we
restricted this assessment to four major regions, each of which has been further delineated into
two to five areas. The basis for the regional delineation is presented in section II.C.
B. FACTORS THAT INFLUENCE SENSITIVITY TO ACIDIC DEPOSITION
The geologic composition of a region can play an important role in influencing the chemistry
and therefore sensitivity of the surface waters to the effects of acidic deposition. Bedrock geology
formed the basis for a national map of surface water sensitivity (Norton et al. 1982) and has been

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Table II.A.1. Regions of North America with large populations of low-ANC surface waters

Predominant


1985 Wet S
1985 Wet NO,



Low-ANC
Sample
Population
deposition
(kg S ha1 yr')
deposition
(kg N ha' yr')
Risk

Region
Resource
Size
Estimate
Evaluation"
Selected References
Adirondack Mtns., NY
Lakes and
Streams
155
1,469a
1,290°
8
4-5
High
Linthurst et al. (1986)
Kretzer et al. (1989)
Catskill Mtns., NY
Streams
66
160
8-10
4
High
Stoddard and Murdoch (1991)
New England
Lakes and
Streams
608
5,800°
4-8
2-3
Moderate
Linthurst et al. (1986); Kahl et al.
(1991)
Appalachians
Streams
344
450*






NA
2,000'
8-10
2-5
High
Kaufmann et al. (1988); Cosby
et al. (1991)

Lakes
94
260c



Elwood et al. (1991)
Florida
Lakes
148
2.100c
5
2
Moderate
Linthurst et al. (1986), Pollman
and Canfield (1991)
Upper Midwest
Lakes
587
8,500°
4-7
2-3
Moderate
Linthurst et al (1986), Glass et al.






(1986); Nichols and McRoberts
(1986)
Eastern Canada
Lakes and
Streams
7400
700,0009
5-10
2-5
High
Jeffries et al. (1986); Jeffries
(1991)
Western Canada
Lakes and
Streams
—
NAg
1-3
< 1
Moderate
—
Alaska
Lakes
992 3
.000,000"
1
< 1
Low
Mauer and Woods (1987); Eilers
et al. (1993)
Western U.S.
Lakes and
Streams
720
10,400'
1-3
1-2
High
Landers et al. (1987)
•	Sisterson et al. (1990)	' Kaufmann et al. (1988)
b Risk based on subjective assessment weighing factors such as s Jeffries (1991)
evidence of acidification, current deposition, trends in	h Alaska Department of Environmental Conservation 305b Report
deposition, and sensitivity of resources	1 Landers et al. (1987)
c Landers et al. (1986) - lakes > 4 ha
" Kretzer et al. (1989) - lakes > 1 ha
*	Virginia Trout Streams

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used in numerous acidification studies of more limited extent (e.g. Bricker and Rice 1989, Dise
1984, Gibson et al. 1983). Analysis of bedrock composition continues to be an important element
for assessing sensitivity of surface waters in mountainous regions (e.g. Stauffer 1990, Stauffer and
Whittchen 1991, Vertucci and Eilers, In Press).
However, the presence of large populations of acidic and low-ANC lakes and streams in
regions such as Florida that are underlain by calcareous bedrock illustrate that if the surface
waters are isolated from highly weatherable minerals, acid-base status is not controlled by
bedrock geology. Many Karst lakes in northern Florida are situated in highly weathered marine
sands that are capable of providing comparatively little neutralization of acidic inputs. For lakes
located above the calcareous bedrock in areas with minimal hydrologic connection with the
Floridan aquifer, the surface waters can be acidic despite groundwaters saturated in carbonate
minerals. Conversely, where calcareous soils have been deposited over resistant bedrock such
as granite, lakes and streams draining such soils are predominantly alkaline. This situation
occurs in northeastern Wisconsin, for example, where a glacial lobe containing calcareous till is
overlain' on Precambrian bedrock. We observe from these examples that both soil and bedrock
composition may exert strong influence on surface water acid-base chemistry, and therefore are
important factors to be considered in defining a sensitive region.
The third principal factor now recognized as critical in contributing to sensitive aquatic
resources is watershed hydrology. The movement of water through the soils, into the lake or
stream, and the interchange with the soils and sediments regulate the type and degree of
response to acidic inputs. Lakes in the same physiographic setting can have radically different
sensitivities to acidic deposition depending on the relative contributions of near-surface drainage
water and deeper groundwater contributions (Eilers et al. 1983, Chen et al. 1984, Driscoll et al.
1991). The movement of water through natural conduits in peat can circumvent hydrologic
routing through wetlands (Gjessing 1992). Even acidic deposition that may not pass through the
watershed, but instead falls as precipitation directly on the lake surface, may eventually be

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neutralized by in-lake reduction processes (Baker and Brezonik 1988). Natural hydrologic events
also radically alter sensitivity to acidification by bypassing normal neutralization processes during
snowmelt or changing flowpaths during extended droughts (Webster et al. 1990). The importance
of hydrologic factors in influencing the acid-base chemistry of surface waters across the United
States was reinforced by Newell et al. (1993), who identified hydrology as a key component
associated with changes in the acid-base chemistry of lakes included in EPA's Long Term
Monitoring Program.
C. REGIONAL DELINEATION
We have identified four regions for inclusion in this assessment: western and upper
midwestern United States, Florida, and southeastern Canada. The information available for
delineating these regions is based almost entirely on data derived from surveys of lakes. This
does not imply that streams in some of these regions may not also be sensitive to acidic
deposition. Rather, it merely reflects the paucity of synoptic stream data suitable for this regional
delineation. Snowmelt chemistry was collected from stream sites in the Sierra Nevada during
1993 (J. Stoddard, pers. comm.), but these data are not yet available. Some stream chemistry
data were collected in the Northern Highlands of Florida by the National Stream Survey
(Kaufmann et al. 1988). Acidic streams in this area, however, were very dilute, weakly-acidic, CP-
dominated systems. The majority of sulfur deposition to the stream watersheds is retailed in
watershed soils, and the mean SO/" concentration was only 22 ^eq LThe very low base cation
concentrations, and possible sodium retention (neutral salt effect) likely contribute to their acidity
(Sullivan 1990). Reviews of data bases in the three U.S. regions we have selected for study
(Baker et al. 1990, 1991) have shown that accurate spatial representations of areas with low ANC
surface waters are provided by data from EPA's synoptic lake surveys (Linthurst et al. 1986,
Landers et al. 1987). We have therefore relied heavily on the data from the Eastern and Western

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Lake Surveys to delineate the regions for analysis, supplementing the survey data with data from
other studies where appropriate.
1. West
The areas containing low-ANC lakes in the West are confined primarily to the higher elevation
mountainous regions, most of which have been glaciated. The Western Lake Survey (WLS,
Landers et al. 1987) sampled some lakes at lower elevation, particularly in the Puget Lowlands.
These systems were among the lakes highest in DOC in the survey, with a median concentration
of 3.5 mg C L'\ thus reflecting the importance of organic acidity in these systems. The few lakes
sampled on the Olympic Peninsula were moderately high-ANC systems and do not warrant
inclusion in the analysis presented here.
We sought to consolidate the study areas in the West into major lake populations located
within similar geomorphic units (Figure 11.0.1). In this way, surface waters with common geologic
features could be logically grouped together into regional airsheds. Minor non-contiguous lake
populations were excluded from the subpopulations of interest, particularly if the population
estimate for the number of lakes totalled < 200 (< 2% of the WLS total population estimate) or
the estimated number of lakes with ANC < 50 /ieq L"1 was < 25. We examined the five
subregions in the West (4A-4E) and identified those geomorphic units that could be best
combined into regional populations of acid-sensitive lakes. The resulting populations account for.
the majority of sensitive lakes in the West, representing over 88% of the lakes in the WLS that had
ANC < 50 ,ueq L'\
The regional delineation for the West based on this approach resulted in defining the
following subpopulations for consideration in the analyses conducted for this report:
• Sierra Nevada. CA: This area includes the Sierra Nevada Mountain Range as defined in
Landers et al, (1987), which contains the most dilute and potentially-sensitive drainage
lakes in the United States (Melack and Stoddard 1991). Other lake groups in subregion

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Olympic
Uplift
Puget
Lowlands
Klamath
Mountains
Oregon
Cascades
CA * N V
North	4Q
Washington	Selklrk	Cabmg(
Mountains Mountains
f—.-M
Middle j(* •jfN
Washington \ V
Cascades I ^
Wenatchee j
Mountains	l\ V
"South I	1
Washington [ water
Cascades	 WA : Mtns.
River
Lewis Range
Salmon River
Mountains
Sawtooth
Mountains
OR ! 10
(California ^
Cascades T~ \
4A
Sierra
Nevada
Bitter root	4Q
Range \
/V Anaconda-Pintlar
fy~i Mountains, Beartooth
y Uplift
MT
. 		, Bighorn
_j — riQ/'X/*/	v. \ Uplift
]V?2	Yellowstone
fi>°>	/
I Ventre '	/ .
/ Uplift n '	/
Wind	/
River Park
Range Range,
WY _ /__ J*s>
T
Medicine Bow
Range
Wasatch
Mountains
<2- /TJ'3
Arch
San Juan
Mountains
1 UT
Sawatch Uplift
Sangre De Cristo
Uplift
\
\
\i
Figure II.C.1. Major geomorphic units and locations of lakes sampled in the Western Lake
Survey (Source: Landers et al. 1987). Those areas included in this analysis of
sensitive lake resources are shaded.

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4A of the WLS, which we do not include in the Sierra Nevada subregion discussed here,
included those in the Klamath Mountains and the California Cascades. The Klamath
Mountain group.had an estimated population size of 153 lakes with only 5 lakes (3%)
having ANC < 50 ^eq L"1. The California Cascades was grouped together with the
remainder of the Cascade Range for our analyses.
•	Cascades. CA, OR, WA: Subregion 4B from the WLS was modified by excluding two
groups of lakes and adding another set. Lakes in the Puget Lowlands were excluded on
the basis of their very different physiographic setting, relatively high DOC values, and
likelihood of greater watershed disturbance. Lakes in the Olympic Peninsula were
excluded because of their high ANC derived from weathering of sedimentary bedrock.
Lakes in the California Cascades (part of subregion 4A of the WLS) were joined with the
remainder of Cascade lakes on the basis of similar geology (andesite and andesitic
basalt).
•	Idaho Batholith. ID. MT: Subregion 4C was modified by excluding lakes in the Blue
Mountains and Wallowa Mountains of Oregon and most of the lakes in Montana. The
lakes that were retained were those associated with the Idaho Batholith. The major
omission using this delineation was loss of lakes in the Lewis Range, MT (N=553).
Unfortunately, only 15 lakes were sampled in the Lewis Range during the WLS, making
the estimates for this area somewhat imprecise.
•	Northwest Wyoming: This area encompasses most of WLS subregion 4D, excluding the
Wasatch Mountains and Uinta Range, UT, located to the south and the Bighorn Mountains
in northcentral Wyoming. The Wasatch Mountains have relatively few lakes; the Uinta
Range has an estimated 20 lakes with ANC < 50 p.eq L'1 and a total estimated
population of over 500 lakes (Landers et al. 1987). The Uinta Range is located ~200 km
from other Wyoming mountain ranges and could be addressed as a distinct
subpopulation at a later date.
•	Colorado Rockies: This diverse grouping of ranges is nearly identical to WLS subregion
4E. Although this area contains the lowest number and percentage of sensitive lakes in
the West, it is also downwind of major emission sources, thus elevating the concern for
this region beyond what would be expected on the basis of lake chemistry alone. The
lakes in the Colorado Rockies share some common features such as a high proportion of
watersheds with exposed bedrock: however the geological diversity of this region warrants
re-examination of separate subpopulations such as the Front Range.
2. Upper Midwest
Several regional lake surveys were conducted across the upper midwestern states of
Minnesota, Wisconsin, and Michigan (Glass and Loucks 1986, Nichols and McRoberts 1986,
Linthurst et al. 1986). All surveys showed similar spatial patterns, whereby the most acidic and
lowest ANC lakes were located in Michigan and the highest ANC lakes were generally located in
Minnesota. Low-ANC lakes (< 50 fieq L"1) are relatively uncommon in Minnesota, and

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therefore the ELS subregions 2A and 2B (using the revised subregional delineation of the Upper
Midwest; cf. Cook and Jager 1991; Figure II.C.2) were omitted from this assessment. This
reduction in geographic extent for the Upper Midwest resulted in an omission of an estimated 210
lakes in the region with ANC < 50 fieq L1 (16%) while eliminating virtually no acidic lakes.
Furthermore, most of the low-ANC lakes in Minnesota have high concentrations of DOC that
would buffer mineral acid inputs. The median DOC concentration for low-ANC lakes (< 50
fieq L'1) measured in this subregion of the ELS was 14.5 mg L'\ with only one sample lake with a
DOC < 3 mg L'1 (Linthurst et al. 1986).
Figure II.C.2. Correspondence between original ELS-I subregions for the Upper Midwest (A)
(Linthurst et al. 1986) and a revised delineation of Region 2 (B) prepared by Eilers
and Bernert (unpublished, map reproduced in Cook and Jager 1991).

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3.	Northern Florida
With few exceptions, low-ANC lakes in Florida are located on ridges extending across the
Panhandle and in the central peninsula (Linthurst et al. 1986, Canfield et al. 1983, Pollman and
Canfield 1991). There are two principal populations ot low ANC lakes in the peninsula, the Trail
Ridge east of Gainesville and the Central Ridge near Orlando. The northern boundary for the
Florida lake district used here is similar to that used for subregion 3B in the ELS (Linthurst et al.
1986) with the exception that lakes in the Okefenokee Swamp of Georgia are excluded. These
high-DOC systems (e.g., > 30 mg'C L ') are naturally acidic and their chemistry is dominated by
organic acids. The southern boundary of the Florida region was adjusted northward to latitude
29° N to exclude additional lakes that have been heavily modified by land use activities. Many of
these lakes in the Central Ridge were either moderate to high DOC systems or apparently have
been highly impacted by agricultural practices in the area. The major ion contributions from
agricultural sources confound inferences regarding potential impacts from acidic deposition (cf.
discussion by Pollman and Canfield [1991]). The two lake districts for northern Florida
represented in this assessment (Figure II.C.3) comprise an estimated 81% of the low-ANC lakes
(< 50 ,aeq L'1) in Florida sampled in the ELS-I.
4.	Southeastern Canada
Southeastern Canada represents an area comparable in size to the eastern United States
and contains about 700,000 lakes of which as many as 12,000 may be acidic (Jeffries 1991). The
area represented by this assessment is that used by Jeffries (1991) and includes lakes from six
provinces (Figure II.C.4). Unlike the data from EPA's National Lake Survey, the data available for
assessing the status of Canadian waters is derived from a collection of independent studies that
were not linked to a statistical frame. Nevertheless, because of the large number of samples
(> 7400), Jeffries (1991) concluded that the collective bias of these surveys is acceptably small.

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Panhandle Lakes
Northcentral Peninsula Lakes
too • 200
> 200
lilM
111
III
Figure II.C.3. Location of lakes sampled in ELS-I (Linthurst et al. 1986) and used in this
subregional analysis of sensitive lake populations. (Source: Pollman and Canfield
1991)

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LAB
NW ONT
QUE
NF
NB
Figure II.C.4.
Subregions of eastern Canada used by Jeffries (1991) in spatial analysis of lake
chemistry. (Source: Jeffries 1991)

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III. APPROACH
A. ACID SPECIES/PRECURSORS
Areas of the United States and Canada having sensitive surface waters are .identified and
described. For each area having low ANC surface waters, regional characteristics are discussed
that influence surface water sensitivity to acidic deposition. Certain regional characteristics may
increase or decrease the sensitivity of surface waters beyond what would be expected based
solely on ANC. For example, organic acids, aluminum, soils, hydrology, deposition history, and
land use can all modify a determination of sensitivity to acidic deposition and can modify the
effects of acidic deposition on aquatic resources.
For the regions identified as having sensitive or critically sensitive aquatic resources, relevant
information has been compiled and evaluated regarding the relationship between deposition
loading (N and S) and the estimated (or expected) extent, magnitude, and timing of aquatic
effects. The general approach we have employed for this task involves a "weight of evidence"
evaluation of the relationships between deposition and effects, as followed by NAPAP in the
Integrated Assessment (NAPAP 1991). This evaluation includes a number of lines of evidence, as
outlined below. Further details regarding the application of these assessment methods are
provided in Appendix A.
Six types of evidence were used in the Integrated Assessment (NAPAP 1991) to assess the
sensitivity of aquatic resources to changes in deposition magnitude and timing:
1.	watershed models that project or hindcast chemical changes in response to changes in
sulfate deposition (particularly the MAGIC model)
2.	biological response models linked to the outputs from the watershed chemistry models
3.	inferences from current chemistry in relation to current levels of deposition
4.	trend analyses based on comparing recent and past measurements of chemistry and
fishery status during the past one or two decades in regions that have experienced large
recent changes in acidic deposition
5.	paleolimnological reconstructions of water chemistry using fossil remains of algae
deposited in lake sediments

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6. results from watershed or lake acidification/deacidification experiments
The assessment provided in the following sections includes, where possible, evaluation of
each of the above-listed analysis approaches. The principal watershed models included in the
assessment are the MAGIC model for drainage systems and the Integrated Alkalinity Generation
(IAG) model (Baker et al. 1986, Baker and Brezonik 1988) for seepage systems.
In Section IV of this report, we present a regional assessment of the known or perceived
relationships between S and N deposition and surface water acidification. Each of the four
selected study regions is examined in detail: West, Upper Midwest, Northern Florida, and Eastern
Canada. These regions were selected because of the sensitivity of their aquatic systems to
potential effects of acidic deposition, not because of the availability of appropriate experimental or
measurement data with which to quantify deposition/effects relationships. In fact, the region that
exhibits the greatest sensitivity (West) is most deficient in terms of available data. We have
therefore included, in addition to the four regional assessment sections of the report (IV.A through
IV.D), an additional section (IV.E) that summarizes deposition/effects relationships in the
Adirondack Mountains, the most extensively studied region in the United States. Data from the
Adirondack region are used in a comparative fashion, to augment our treatment of the resources
within the selected study regions. Such comparisons are necessary because of the paucity of
appropriate data for the study regions of concern.
1. Sulfur
Several watershed processes control the extent of ANC consumption and rate of cation
leaching from soils to drainage waters as water moves through undisturbed terrestrial systems.
Of particular importance is the concentration of anions in solution. Naturally-occurring organic
acid anions, produced in upper soil horizons, normally precipitate out of solution as drainage
water percolates through lower mineral soil horizons. Soil acidification processes reach an
equilibrium with acid neutralization processes (e.g., weathering) at some depth in the mineral soil

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(Turner et al. 1990). Drainage waters below this depth generally have high ANC. The addition of
strong acid anions from atmospheric deposition allows the natural soil acidification and cation
leaching processes to occur at greater depths in the soil profile, thereby allowing water rich in
mobile anions to emerge from mineral soil horizons. If these anions are charge balanced by
hydrogen and/or aluminum cations, the water will have low pH and could be toxic to aquatic
biota. Thus, the mobility of anions within the terrestrial system is a major factor controlling the
extent of surface water acidification.
Sulfate is the most important anion, on a quantitative basis, in acidic deposition in most parts
of the United States. Consequently, sulfate and the controls on its inputs and processing have
received the greatest scientific and policy attention to date (Turner et al. 1990). Virtually all of
NAPAP's major aquatic modeling and integration efforts focussed predominantly on the potential
effects of sulfur deposition (e.g., Church et al. 1989, Turner et al. 1990, Baker et al. 1990b,
Sullivan et al. 1990a). The response of watersheds to sulfur inputs, particularly chronic effects on
surface water quality, are now reasonably well understood. This understanding has been
developed largely through the efforts of three large multidisciplinary research efforts: the
Norwegian SNSF program (Acid Precipitation Effects on Forests and Fish, 1972-1980), NAPAP
(1980-1990), and the British-Scandinavian Surface Water Acidification Program (SWAP 1984-
1990).
2. Nitrogen
The second important acid anion found in acidic deposition, in addition to sulfate, is nitrate.
Nitrate (and also ammonium, which can be converted to nitrate within the watershed) has the
potential to acidify drainage waters and leach potentially toxic Al from watershed soils. In most
watersheds, however, nitrogen is limiting for plant growth, and therefore most nitrogen inputs are
quickly incorporated into biomass as organic nitrogen with little leaching of nitrate into surface
waters. A large amount of research has been conducted in recent years on nitrogen processing

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mechanisms and consequent forest effects, mainly in Europe (Sullivan 1993). In addition, a
smaller N research effort has been directed at investigating effects of nitrogen deposition on
aquatic ecosystems. For the most part, measurements of nitrogen in lakes and streams have
been treated as outputs of terrestrial systems. However, concern has been expressed regarding
the role of N03' in acidification of surface waters, particularly during hydrologic episodes, the role
of N03" in the long-term acidification process, the contribution of NH4+ from agricultural sources to
surface water acidification, and the potential for anthropogenic N deposition to stimulate
eutrophication of freshwaters and estuaries.
Nitrate in snowmelt runoff has been recognized for some time as an important component of
biological damage resulting from atmospheric deposition (cf. Wigington et al. 1990). Nitrate is the
principal acid anion in snowmelt in many areas of northern Europe and the northeastern United
States. Selective elutriation of N03 from the snowpack can result in early spring runoff having
concentrations substantially greater than the average snowpack concentrations. The biological
response to acidic runoff is similar, regardless of whether the predominant acid anion is N03" or
S042', assuming concentrations of other ions, including inorganic aluminum (Al;), are the same.
However, there is some concern that N03 may be more effective than SO/ in mobilizing Al,
(Driscoll et al. 1991).
Concern for chronically elevated nitrate concentrations in aquatic ecosystems received
considerably greater attention in 1988 following the publication of the resurvey of Norwegian lakes
(SFT 1987). Over 1000 lakes, 305 of which were originally sampled in 1974/75 (Wright and
Henriksen 1978), were sampled again in 1986 (SFT 1987, Henriksen and Brakke 1988). Even
though the average S042' concentration declined in the lakes, the pH remained virtually
unchanged because of increased N03 and decreased base cation concentrations. In the
southern portions of Norway, N03" concentrations in the lakes doubled between 1974/75 and
1986, reaching county-wide average concentrations as high as 14,ueq L'1 in Rogaland County
and up to 50 /*eq L'1 in individual lakes (Henriksen and Brakke 1988). An analysis of fisheries in

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the study lakes showed an increase in the number of fishless lakes, perhaps attributable to the
concomitant increase in labile Al and decrease in (Ca2+ + Mg2+) (SFT 1988). Analysis of selected
lakes and streams with longer-term records also showed increases in nitrate concentrations,
providing additional evidence for an increasing trend in N03'. Although S042" remained the
dominant anion in most systems, the ratio of N03/(N03 + S042") reached 0.54 on an equivalent
basis in some lakes and rivers in southwestern Norway (Henriksen and Brakke 1988). These
authors summarized the ratio of N03"/(N03 + S042') for many acidified waters in Europe and
North America, illustrating that the relative importance of N03' in acidified surface waters is
substantial, particularly in central Europe.
Until quite recently, atmospheric deposition of nitrogen has not been considered detrimental
to either terrestrial or aquatic resources. Because most atmospherically-deposited nitrogen is
strongly retained within terrestrial systems, atmospheric inputs of nitrogen have been viewed as
fertilizing agents, with little or no nitrogen moving from terrestrial compartments into drainage
waters. More recently, however, biogeochemical nitrogen cycling has become the focus of
numerous studies at the forest ecosystem level. It has become increasingly apparent that, under
certain circumstances, atmospherically-deposited nitrogen can exceed the capacity of forest
ecosystems to take up nitrogen. This nitrogen saturation can lead to base cation depletion, soil
acidification, and leaching of N03 from soils to surface waters. Aber et al. (1989) provided a
conceptual model of the changes that occur within the terrestrial system under increasing loads
of atmospheric nitrogen. Stoddard (1994) described the aquatic equivalents of the stages
identified by Aber et al. (1989), and outlined key characteristics of those stages as they influence
seasonal and long-term aquatic nitrogen dynamics.
Nitrate concentrations in surface waters exhibit a strong seasonality; N03 is typically elevated
during late winter and spring, particularly during periods of snowmelt, and reduced to low or non-
detectable levels throughout summer and fall. This can be attributed to seasonal growth patterns
of forest vegetation. Vegetation growth is reduced or stopped entirely during winter months, and

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microbial assimilation of N is also reduced during this season. Spring snowmelt can act to flush
into lakes and streams nitrogen that was deposited in the snowpack from atmospheric deposition
or nitrogen mineralized within the soil during winter.
A number of factors can be involved in controlling the loss of nitrogen from a forested
watershed, including atmospheric inputs, forest stand age and condition, soil nitrogen pools, and
flowpaths of percolation and melt-water within the catchment. The stages of N loss from the
watershed, as described by Stoddard (1994), are depicted in Figures III.A.1 to III.A.4. Changes in
both the seasonal and long-term patterns in surface water N03 concentrations reflect changes
that are occurring within the watershed in nitrogen cycling and the degree of nitrogen saturation.
At Stage 0, N03' concentrations in drainage waters are very low throughout most of the year, and
increase to measurable concentrations typically only during snowmelt or spring rainfall
hydrological events (Figure III.A.1). The loss of N in runoff is short-lived and small in magnitude.
This is viewed as the "natural" pattern. At Stage 1, the natural pattern is amplified; spring
concentrations of N03 in surface waters reach relatively high concentrations and the seasonal
onset of N limitation is delayed (Figure III.A.2). In Stage 2, N begins to percolate beneath the
rooting zone of the soil, resulting in elevated groundwater concentrations of N03. Seasonality is
damped because baseflow concentrations of N03' are high (Figure III.A.3). In Stage 3, the
watershed becomes a source, rather than a sink, for atmospheric N. The combined inputs of N
from deposition, mineralization, and nitrification produce concentrations of N03' in drainage water
of Stage 3 watersheds that can be higher than deposition (Figure ill.A.4, Stoddard 1994).
Except in cases of excessive nitrogen saturation, the effects of nitrogen deposition on surface
waters are expected to be primarily episodic in nature. Unfortunately, data required to make
regional assessments of episodic effects are generally not available. Such data need to be
collected on an intensive schedule and must include sample periods during late winter and early
spring when snowmelt often causes the most severe nitrogen-driven episodes of surface water
acidification. Sampling during this time of year is more difficult and expensive than during the

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DEPOSmON
wot dry
NO- nh;
Stage 0
WINTER/SPRING
SUMMER/FALL
DEPOSITION
(Hy «nt
K no-
(tear tall)

Black Pond, Adirondacks, New York
1985	1986	1987
Figure III.A.1. Schematic representation of the nitrogen cycle at Stage 0 of watershed nitrogen
saturation, with an example pattern of N03' concentration in surface water from
Black Pond in the Adirondack Mountains, NY. Nitrogen transformations are
dominated by plant uptake and microbial assimilation. Only small amounts of NOa
leach out of the watershed, primarily during snowmelt. (Source: Stoddard 1994)

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Stage 1
SUMMER/FALL
WINTER/SPRING
DEPOSITION
mh! no-
NO- NH4
uotaw

Constable Lake, Adirondacks, New York
40
1986
1985
1987
Figure III.A.2. Schematic representation of the nitrogen cycle at Stage 1 of watershed nitrogen
saturation, with an example pattern of N03" concentration in surface water from
Constable Pond in the Adirondack Mountains, NY. As compared to Stage 0, this
stage is characterized by delayed onset of N limitation during spring and larger
peaks of N03' in runoff. (Source: Stoddard 1994)

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Stage 2
SUMMER/FALL
WINTER/SPRING
no; NH
Fernow Experimental Forest, West Virginia
100	:	^	i	
~Z. 20 -
o				1 1 			
1985	1986	1987
Figure III.A.3. Schematic representation of the nitrogen cycle at Stage 2 of watershed nitrogen
saturation, with an example pattern of N03" concentration in surface water from
Watershed #4 at Fernow Experimental Forest, West Virginia. Uptake of N by
plants and microbes is reduced, as compared with Stage 1, resulting in loss of
NO, to streams during winter and spring and to groundwater during the growing
season. (Source: Stoddard 1994)

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Stage 3
DEPOSITION
wm
NQi
WINTER/SPRING
SUMMER/FALL
DEPOSITION
dry wet
*K no.
(litter fall)
T
No°
2.2
uotake
(jnineralaati
tneraltzBtiory
ftrfficailo
(pitrtficatio
denitrificatiofv
groundwater
Dicke Bramke, Harzburg Mountains, Germany
300

1987
1986
1985
Figure III.A.4. Schematic representation of the nitrogen cycle at Stage 3 of watershed nitrogen
saturation, with an example pattern of N03" concentration in surface water from the
Dicke Bramke stream site in Germany. At Stage 3 potential N sinks in the
watershed are saturated and all inputs, as well as mineralized N, are exported from
the watershed. (Source: Stoddard 1994)

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more common summer/fall sampling seasons. Sampling during snowmelt can be particularly
difficult in the high mountains of the West, when study sites are often inaccessible, and when
motorized transport (e.g., via snowmobile) is often not allowed due to wilderness restrictions.
B. TEMPORAL RESPONSE
1. Chronic Acidification
The U.S. EPA's National Lake Survey (NLS), conducted in 1984 and 1985, provided the most
comprehensive data base on the acid-base chemistry of lakewaters in areascof the United States
potentially susceptible to the effects of acidic deposition. This synoptic survey was conducted
during the autumn "index period," during which time lakewater chemistry typically exhibits low
temporal and spatial variability. Although autumn is an ideal time for surveying lakewater
chemistry in terms of minimizing variability, lakewater samples collected during autumn provide
little relevant data on the dynamics or importance of nitrogen in most aquatic systems. Nitrate
concentrations in lakewater will be elevated during the autumn season only in lakes having
watershed that exhibit fairly advanced symptoms of nitrogen saturation (e.g., Figure III.A.3,
Stoddard 1994). It is therefore not surprising that results of both the Eastern and Western Lake
Surveys suggested that N03 is of only minor importance compared to S042 as an acid anion in
lakewaters in this country. For example, the median value of the ratio of lakewater[N03'] to [S042'
+ N03'] in both Florida and the Upper Midwest was only 0.01 (Stoddard 1994). In subregions of
the West, this ratio varied from 0.01 to 0.06 (Landers et al. 1987). Chronic effects from acid
deposition are certainly more likely to be associated with sulfur, rather than nitrogen, in these
regions. In addition, survey data with which to evaluate the (largely episodic) effects that might
be associated with nitrogen deposition were not collected in the NLS.
Chronic acidification from nitrogen deposition is not of widespread occurrence in the United
States. Highest chronic N03' concentrations have been found in the Great Smoky Mountains,
where baseflow N03' concentrations approximate S042' concentrations (Cook et al. in press) and

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N03" concentrations from deep soil lysimeters have been found to be higher than input from
deposition (Johnson et al. 1991, Stoddard 1994). Chronically elevated N03" in the Great Smoky
Mountains appears to be the exception, rather than the rule, for systems in the United States. In
contrast, many European sites show chronically elevated N03' concentrations in soil waters and
drainage waters (Hauhs et al. 1989, Henriksen and Brakke 1988). This is undoubtedly due to the
much higher levels of atmospheric N deposition that are found throughout much of Europe.
A variety of model approaches are being used for estimating the long-term (chronic) critical
loads of S and N to surface waters. They range from simple empirical calculations to complex
dynamic models. Steady state models can be useful to derive long-term critical loads for sulfur,
and potentially for nitrogen. They only include processes that influence acid production and
consumption over long periods of time, such as mineral weathering and net uptake. An
assumption in the application of steady state models is that dynamic processes are not important
for the assessment of long-term critical loads. Dynamic models include evaluation of the time
period required to reach critical criteria values. Thus, processes such as cation exchange,
nitrogen mineralization/immobilization and sulfate adsorption/desorption are often included in the
dynamic approaches (deVries and Kros 1991). Although steady state models will provide
estimates of the final emission or deposition amounts required to achieve a steady state condition
over an infinite time period, dynamic models are needed for an assessment of the temporal
evolution of the acidification process.
A large number of both steady state and dynamic models have been developed and are
being used in the assessment of acidic deposition effects and the development of acidic
deposition standards (see e.g., reviews of Eary et al. 1989, Thornton et al. 1990, deVries and Kros
1991). The model most commonly used for policy projections in the United States (e.g. Church
et al. 1989, NAPAP 1991) is the Model of Acidification of Groundwater in Catchments (MAGIC)
(Cosby et al. 1985a,b). The MAGIC model has been described in the NAPAP State of
Science/Technology Report on modeling methods (Thornton et al. 1990), and numerous model

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projections of future scenarios were presented in the Integrated Assessment (NAPAP 1991). For
seepage lakes, which are common in Florida and in parts of the Upper Midwest, the principal
model used during the NAPAP research program was the Internal Alkalinity Generation (IAG)
model (Baker et al. 1986, Baker and Brezonik 1988, Pollman and Sweets 1990). The IAG model
was developed for seepage lakes, and is based on alkalinity/electroneutrality principles and
input/output budgets analogous to those used in nutrient loading models (e.g., Vollenweider
1975). MAGIC and IAG, as well as static models and empirical approaches, are further discussed
in Appendix A.
Diverse data are available from a variety of sources with which to quantify the watershed
acidification response, as well as recovery from acidification. Such data shed light on the
sensitivity of various kinds of watershed systems to changes in acidic deposition.
Intercomparisons among the various studies that have been conducted are complicated by
different relative watershed sensitivities, sulfur deposition loading rates (and changes in those
rates), relative importance of nitrogen leaching and nitrogen saturation, temporal considerations,
and natural (especially climatic) variability. In addition, these quantitative data have been
generated in vastly different ways, including monitoring, whole-watershed or whole-lake
acidification, whole-watershed acid exclusion, paleolimnology, and modeling. The only way in
which different approaches can be compared on a quantitative basis is by normalizing surface
water response as a fraction of the change in sulfate concentration (or S042' + N03 concentration
where N03 is also important). The principal ions that change in direct response to changes in
(S042" + N03) concentration are ANC (which can be expressed as [HC03 - H+]), base cations
(CB), inorganic aluminum (Al,), and organic acid anions (A). The proportional changes in [HC03 -
H+], [Alj], [CB], and [A ] should sum to 1.0 in order to satisfy the electroneutrality condition. For
aquatic systems that are relatively insensitive to acidic deposition, A [CB] approximates A [S042' +
N03], or the F-factor (Henriksen 1982) approximately equals 1.0:

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A [SO,2" * NO3]
and changes in other constituents are insignificant. Where acidification occurs in response to
acidic deposition, changes in [HC03 - H+] and/or [Al,] comprise an appreciable percentage of the
overall surface water response and therefore the F-factor is less than 1.0 (Sullivan 1990). The F-
factor is important in evaluating criteria for establishing acid deposition standards because it
provides the quantitative linkage between inputs of acid anions (e.g., S04z', N03) and effects on
surface water chemistry.
2. Episodic Acidification
The acid-base chemistry of surface waters typically exhibits substantial intra- and interannual
variability. Seasonal variability in the concentration of key chemical parameters often varies by
more than the amount of acidification that might occur in response to acidic deposition. Such
variability makes quantification of acidification and recovery responses difficult, and also
complicates attempts to evaluate sensitivity to acidification based solely on "index" chemistry.
Lakes and streams exhibit short-term episodic decreases in ANC, and often also pH, usually
in response to hydrological events, such as snowmelt or rainfall. Periods of episodic acidification
may last for hours to weeks, and sometimes result in depletion of ANC to negative values with
concurrent increases in potentially-toxic inorganic Al in solution. The occurrence and magnitude
of episodic acidification was summarized in the review of Wigington et al. (1990). Episodes are
generally accompanied by changes in at least two or more of the following chemical parameters:
ANC, pH, base cations (CB), S042', N03", Aln+, organic acid anions (A), and dissolved organic
carbon (DOC). These changes in chemistry can adversely impact biota, particularly when
changes involve pH, inorganic Al (Al;), and/or calcium (Baker et al. 1990a). Aquatic biota vary
greatly with respect to their sensitivity to episodic decreases in pH and increases in Al; in waters

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having low Ca2+ concentration, and it is difficult to classify chemical episodes based on potential
biological effects. Baker et al. (1990a) concluded, however, that episodes are most likely to
impact biota if the episode occurs in waters with baseline (pre-episode) pH above 5.5 and
minimum pH during the episode of less than 5.0. In addition, for episodes that occur in systems
that are chronically acidic or nearly so, the increase in acidity during the episode may be
biologically significant, particularly when it is accompanied by increased concentrations of Al;
(Baker et al. 1990a).
Episodic acidification is nearly ubiquitous in drainage waters. Lakes and streams that have
been studied throughout the United States, Canada, and Europe nearly all experience loss of
ANC during hydrologic events (Wigington et al. 1990). Chemical changes during episodes are
controlled by a number of natural processes, including dilution of base cation concentrations,
nitrification, flushing of organic acids from terrestrial to aquatic systems, and the neutral salt
effect. Acidic deposition can also contribute to episodic acidification, particularly via enhanced
NOa" leaching. Under some conditions, episodes can also be partially caused by increased SO„2
concentration, although sulfur-driven episodes appear to be the exception, rather than the rule.
There is also the possibility that chronic acidification by acid deposition can pre-condition a
watershed, thereby increasing the severity of episodic acidification.
Most research on episodic processes has been conducted on stream systems, which are
generally more susceptible to such effects than are lakes. Spatial variability can be considerable
in lakes, particularly during snowmelt episodes. Strong vertical and horizontal gradients in
lakewater chemistry often preclude quantification of the magnitude of the effects in lake systems
(Gubala et al. 1991). Because of the logistical difficulties and expense associated with sampling
lakewater chemistry during episodic events in a manner sufficient to characterize these spatial
gradients, few data are available for lakes in the areas of concern for this assessment report.

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IV. REGIONAL ASSESSMENT
The sensitivity to acidification of surface waters in a region is a function of regional deposition
characteristics, surface water chemistry, and watershed factors previously described in Section
MB. The following section attempts to integrate these three elements to provide a qualitative
assessment of watershed sensitivity to acidification and a quantitative assessment of the
magnitude of acidification currently experienced within the study regions. These results are
combined in Section V to provide an assessment of the likely dose/response relationships for the
regions of interest, which is then used to assess the feasibility of adopting one or more acid
deposition standards.
A. WEST
1. Characteristics and Sensitivity to Acid Deposition
In the far West, the sensitive regions form two nearly continuous ranges, the Sierra Nevada in
California and the Cascades starting in California and extending through Washington. The Rocky
Mountains, in contrast, are discontinuous ranges with highly variable geological composition.
Consideration of the Rocky Mountains as a single range for the purpose of evaluating sensitivity
to acidic deposition is merely a convenience. Precise assessments of the sensitivity of Rocky
Mountain resources need to be specific to individual ranges (Turk and Spahr 1991). Portions of
the mountainous West are similar to the highly sensitive areas in Norway where watersheds
contain large areas of exposed bedrock, with little soil or vegetative cover to neutralize acidic
inputs. This is particularly true of alpine regions of the Sierra Nevada, northern Washington
Cascades, the Idaho batholith, and portions of the Rocky Mountains in Wyoming and southern
Rockies of Colorado. However, the percentage of exposed bedrock in a watershed can yield
misleading information on the sensitivity of the resources if the bedrock contains even small
deposits of calcareous minerals or if physical weathering such as that cased by glaciers causes a
high production of base cations within the watershed (Drever and Hurcomb 1986). Consequently,

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even though much of the northern Washington Cascades is alpine, many surface waters draining
glaciers in this area are moderately alkaline.
Topographic relief is also a contributing factor to acidic deposition sensitivity in the West
because the mountainous terrain contributes to major snowmelt events that may cause episodic
pH and ANC depressions. These snowmelt events can last two months and result in multiple
exchanges of the water volume in lakes receiving significant runoff. The short residence times in
the lakes not only contributes to elevated sensitivity to snowmelt events, but also reduces the
relative importance of in-lake alkalinity generation processes. Lakes in the West are generally
short residence time systems, with the exception of a number of lakes in the volcanic portions of
the Cascades where low-ANC seepage lakes are abundant. These seepage lakes share greater
similarity to lakes in the Upper Midwest regarding their sensitivity to acidification than to lakes
elsewhere in the West because of their hydrology and importance of in-lake processes.
Soils in the watersheds of the West, where present, vary considerably from highly alkaline
Mollisols to acidic Spodosols. In the volcanic regions of the West, the soils are typically highly
siliceous and generate little alkalinity. In alpine regions of the West, soils are absent or poorly
developed.
2. Current and Projected Future Deposition
Emission of S02 and N0X in the West peaked in the 1970's and decreased through the
1980's (Placet et al. 1990). Sources of S02 in the West are more diverse than in the East where
the major source is the utility sector. Industrial and "other" categories constitute major sources of
S02 in the West, in addition to utilities. Emissions from mobile sources constitute the largest
contribution of atmospheric NOx in the West, accounting for about one-half of the total
anthropogenic emissions. Natural sources of NOx are also regionally important, and represent
17% of the total nitrogen emissions. Barring significant technological breakthroughs in NOx or
SOx emission control, recent emissions trends suggest that further emission reductions can be

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expected, but that these declines will be both modest and gradual as older, more polluting
industrial sources are removed from operation.
Atmospheric deposition of both sulfur and nitrogen is currently low throughout most portions
of the West (Figure IV.A.1). Annual wet deposition of S and N are generally less than about one-
fourth of the levels observed in the high-deposition portions of the northeastern United States
(Sisterson et al. 1990).
Because of the proximity of well-defined population centers and industrial pollution sources in
the West to individual mountain ranges, it is much more important to evaluate changes in
emissions in the immediate vicinity of sensitive resources. For example, emissions in the Rocky
Mountain states have no effect on resources in the Sierra Nevada and the Cascades, in part
because emissions from these states are generally low and in part because the prevailing wind
direction is from west to east. Precipitation chemistry in these far western ranges is largely
influenced by local emissions in California, Oregon, and Washington, particularly emission
sources to the west (upwind) of sensitive resources. In the Rocky Mountains, deposition
chemistry is influenced by a more complex collection of sources, although recent evidence
suggests that within-state sources' are much more important than long-range sources. In the Mt.
Zirkel Wilderness of northwest Colorado, elevated concentrations of SO/'and N03 in the snow
appear to originate largely from sources in the Yampa Valley, about 75 km to the west (Turk et al.
1992). The conclusion from analysis of deposition in the West is that potential impacts to
sensitive resources appear to be governed primarily by local emissions rather than regional
emissions.
3. Current Surface Water Chemistry
The current chemistry of surface waters in the West is basea almost exclusively on synoptic
data from the Western Lake Survey (Landers et al. 1987) and a small number of more localized
those sources within ~ 100 km of a given mountain range

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max = 1 / .
50%_._4.0
25% L-2-5
min = 0.8
Based on:
NADP/NTN. UAPSP./OEN. MAP3S
CAPMoN. APIOS-C. APIOS-D
REPQ. FADMP
max = 42.7
Annual 1985-1987
Sulfate Deposition
kg ha-1
-1.32.3
.1.28.6
75%... 23.8
5o%_..17.3
25%___14.7
i o%	11.4
min = 5.5
Figure IV.A.1. Color composites of annual wet deposition (kg ha ) of A) SO„2', B) N03\ and
C) NH/ for eastern and western North America for the period 1985 to 1987.
(Source: Sisterson et al. 1990).

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max =11.
75%. 1.5.4
50%_:_3.3
0%.._1.9
min = 0.9
Based on:
NADP/NTN. UAPSP/OEN. MAP3S
CAPMoN APIOS-C. APIOS-D
REPQ.FADMP
Annual 1985-1987
Nitrate Deposition
kg ha-1
J max = 29.0
95%_ j _20.7
90%. 5.18.3
75%... 13.9
^o%	10.9
:s%---9.2
i%.._7.S
min = 5.5
max = 4.2
Based on:
NADP/NTN. UAPSP/OEN. MAP3S
CAPMoN. APIOS-C. APIOS-D
REPQ. FADMP
Annual 1985-1987
Ammonium Deposition
ka ha'1
max = 5.4
95%. 1.4.0
90%J_3.7
75%	3.1
5o%___2.5
25%	2.0
10%	1.4
min = 0.9
Figure IV.A. 1. Continued.

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studies. Comprehensive assessments of lake chemistry in the West (cf. Baker et al. 1990b)
indicate that there are many low-ANC systems, but virtually no acidic waters. The few acidic
waters that have been sampled in the West (e.g., Fern Lake, WY, Eilers et al. 1987 and West Twin
Lake, OR, Eilers and Bernert 1991) are acidic from natural production of sulfur from watershed
sources. It should be noted, however, that the data base available to evaluate the occurrence of
acidic lakes in the West is fairly limited.
The surface water chemistry in areas of interest delineated for the West (see Section II.C.1)
shows that the Sierra Nevada and Cascade Mountains constitute the mountain ranges with the
greatest number of sensitive resources (Tables IV.A. 1 and IV.A.2). Sensitive lakes in the Sierra
Nevada and elsewhere in the West are dilute bicarbonate systems and, unlike many low-ANC
lakes in the East, have very low concentrations of dissolved organic carbon.
Weathering reactions in the West generate a higher percentage of cations other than calcium
in surface waters, as observed in the relative abundance of dissolved magnesium and sodium.
Acid anion concentrations in most western lakes are extremely low in fall samples, but limited
analyses of lake chemistry in spring generally show higher concentrations of N03 and S042" in
both the Sierra Nevada (Williams and Melack 1991b, Melack et al. 1989) and Rocky Mountains
(Reuss et al. In Press). The extremely dilute nature of many western lakes raises concerns
regarding potential increases in acid anions, derived from acidic deposition, during spring
snowmelt: The available data from intensive study sites in the West (e.g., Loch Vale, CO,
Emerald Lake Basin, CA, and the Glacier Lakes Watershed, WY) suggest that episodic depression
of stream pH is more pronounced than for lake systems, yet no systematic regional stream
chemistry data are available in the West with which to assess the regional sensitivity of streams to
acidic deposition or the importance of episodic processes to stream chemistry.
One issue often overlooked in assessments of acidification is the potential contribution of
acid anions other than S0„2" and N03. For example, F" is the dominant acid anion in lakes of the
Sawtooth Mountains of Idaho, with a median concentration of 24 ueq L"1 (Vertucci and Eilers, in

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Table IV.A. 1. Population estimates of water chemistry percentiles for lake populations in the western United States'. The 1st and 5th percentiles (P„
P,) are presented for pH, ANC (ueq L"'), and Ce fceq L1) and the 95th and 99th (P^, PJ percentiles are shown for SO/' («eq L'1) and
NO, (/ a
31 3J £

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Table IV.A.2. Population estimates for the percentage of lakes in the West with ANC and
N03" within defined ranges.3


ANC (uea L'l

NO,"
(ueq L"'l

< 0
< 25
< 50
> 5
> 10
Sierra Nevada
0
8.7
39.3
10.6
1.5
Cascades
0
10,2.
22.4
1.5
0
Idaho Batholith
0
2.0
23.6
4.6
3.9
NW Wyoming13
0
2.3
12.8
22.8
8.9
Colorado Rockies
0
0.9
5.5
9.8
1.8
a Source: Landers et al. 1987
b Excluding Fern Lake (4D3-017) which is a naturally acidic lake
press). The F in solution is attributed to weathering of fluorapatite deposits. Fluoride
contributions have a similar effect on acid-base status1 of surface waters as do watershed sources
of sulfur, which are also widespread throughout the West.
Concentrations of sulfate in western lakes are generally low, but in some cases watershed
sources contribute substantial amounts of S which are evident in lakewater concentrations for
western lakes (Table IV.A.1). Turk and Spahr (1991) presented a conceptual model for expected
sulfate distributions in the West that can be used as an aid in identifying the proportion of
watersheds with significant watershed sources of S. Considering that atmospheric sources can
account for generally < 30 peq L1 of S04 in the West (including the Colorado Rockies), it is clear
that many lakes, particularly in Colorado, receive major, albeit variable, sources of watershed
sulfate (Figure IV.A.2).
Nitrate concentrations are virtually undetectable in most western lakes in the fall (Landers et
al. 1987). However, in some cases, fall nitrate concentrations were surprisingly high (Table
IV.A.1). For example, nearly one fourth of the lakes in NW Wyoming had N03 > 5 ^eq L'1 and

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almost 10% had N03" > 10 fieq L'1 (Table IV.A.2). The Sierra Nevada and Colorado Rockies
subregions also exhibited many lakes with higher N03' concentrations than would be expected for
fall samples. In both areas about 10% of the lakes had N03 concentrations above 5 /*eq L'1
(Table IV.A.2).
In contrast to S, which is generally conservative in the West (Stauffer 1990), N is assimilated
rapidly in most watersheds. Whereas a conceptual model of S distribution in lakes suggests that
most lake populations will exhibit a normal distribution around some positive value (Turk and
Spahr 1991), a companion conceptual model for N presented here suggests that N03'
distributions for undisturbed watersheds should be highly skewed towards zero (Figure IV.A.3).
As N loading exhausts the capability of the watershed to assimilate N03' and NH4+, "leakage" will
be exhibited as an extended regional distribution (Figure IV.A.3). Where watershed disturbance is
severe (e.g., logging, cattle grazing, cropland, urbanization), N03 concentrations in drainage
waters can be substantial. The significant differences between the expected distributions of S
and N are that S is generally conservative, and there are no mineral watershed sources of N, thus
high N concentrations are almost always indicative of anthropogenic sources.2 Distributions of
N03' concentrations in the western lakes are universally low (Figure IV.A.4), but even relatively low
concentrations of N03 may be significant in view of: (1) the low base cation concentrations in
many lakes; (2) trends in regional N emissions which are relatively stable; (3) potential for
continuing N deposition to eventually exhaust natural assimilative capabilities; and (4) the fact that
these distributions are based on a fall survey. Limited data collected during snowmelt indicate
spring concentrations several times higher than data collected in the fall (e.g. Reuss et at., in
press).
Although it is evident from Tables IV.A.1 and IV.A.2 that lakes in the Sierra Nevada,
Cascades, and Idaho are generally lower in ANC that those in Northwest Wyoming and Colorado,
2 An exception occurs where alder trees are common in riparian zones, because alder is
associated with nitrogen-fixing microorganisms.

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SULFATE
SULFATE
Figure IV.A.2.
Sulfate (,ueq L"') distribution in lakes in A) Sierra Nevada, B)
Cascade Mountains, C) Idaho Batholith, D) NW Wyoming, and E)
Colorado Rockies. Data from the Western Lake Survey.

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SULFATE
SULFATE
150
SULFATE
Figure IV.A.2. Continued.

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100
1 2 3 4 5 6 7 8 9 10 11 12 13 14 15
NO3 Concentration (|ieq L"1 )
50
B
1 2 3 4 5 6 7 8 9 10 11 12 13 14 15
NO 3" Concentration (^eq L*1)
Figure IV.A.3. Conceptual model of regional lakewater nitrate distributions in (A) undisturbed
lake populations where the percentage of lakes (Y-axis) with measurable N03"
concentrations (X-axis) is low. As N deposition exceeds the capability for
terrestrial and aquatic systems to assimilate inorganic N, concentrations of N03
in lake populations are expected to show greater dispersion, as shown in (B).

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1 23456789 10 11 12 131415
NITRATE
1100r
B
01 23456789 10 11 12 131415
NITRATE
Figure IV.A.4. Nitrate (ueq L') distribution in lakes in A) Sierra Nevada, B) Cascade Mountains,
C) Idaho Batholith, D) NW Wyoming, and E) Colorado Rockies. Data from the
Western Lake Survey

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600
O
>
o
2
HI
D
O
LU
tt
400'
300-11
200-1
1001
2 3 4 5 0 7 8 9 10 Tl 12 13 14 15
NITRATE
2 3 4 S 6 7 8 9 10 11 12 13 14 15
NITRATE
>
u
z
111
3
o
HI
DC
01 23456789 10 11 12 13 1415
NITRATE
Figure IV.A.4. Continued.

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it also appears that lakes in the central and southern Rockies have considerably higher
concentrations of acid anions. Turk et al. (1992) noted that concentrations of SO,2"and N03" in
the snowpack of northern Colorado are several times greater than in snow collected in southern
Colorado. Nitrate concentrations in the Wyoming and northern Colorado lakes are also high
relative to other sites in the West, and this is noteworthy for two reasons. First, these population
estimates are derived from fall samples when inorganic nitrogen concentrations are expected to
be at minimum levels. Nitrate concentrations in these lakes are expected to be much higher
during spring. Second, there may be an important bias in the WLS data from Wyoming and
Colorado which has caused regional lakewater N03" concentrations to be underestimated. During
the survey, 49 lakes in the statistical frame were not sampled because the lakes were frozen when
the field crews visited the lakes (Landers et al. 1987). Most of the frozen lakes were from
Wyoming (n=21) and Colorado (n=20). Since the frozen lakes were generally higher elevation
sites, it is likeiy that the actual population values for northeast Wyoming and Colorado are higher
in N03' and lower in ANC and CB than previously reoorted. Although the magnitude of this bias
may not have been crjtical to the primary objective of the WLS (to perform regional
characterization of lake populations), it may be important in view of increasing evidence of
elevated deposition in the Rockies (e.g., Turk et al. 1992) and elevated N03" concentrations in
many high-elevation lakes.
Comparison of the chemistry of mountain lakes in Norway and Italy with those in the western
United States shows that lakes in the mountainous western United States are comparable to lakes
in these areas of Europe and Scandinavia (Table IV.A.3). Deposition chemistry in the Alps has
higher pH than in Norway, but this is attributed to higher deposition of NH4+ in Italy (Wathne et al.
1989). Retention of inorganic N in the Norwegian watersheds was estimated at 80%, whereas
only 47% of N was assimilated in the Alps. Wathne et al. (1989) estimated that historical ANC
losses in Norway and the Italian Alps have been 32 ^eq L'1 and 43 ,ueq L'\ respectively. This
model scenario was run assuming an F-factor of 0.2, where F included nitrate and ammonium

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Table IV.A.3. Population estimates tor selected variables in the West compared to survey data from Norway and Italy". All units except pH are In
microequivalents per liter.
Population
Sample
Size

PH


ANC


Ca,+


SO
2

NO,

Min
Mean
Med
Min
Mean Med
Min
Mean
Med
Mean
Med Max
Mean
Med
Max
Sierra Nevada
114
5.84
7.02
6.95
13
105
66
7
79
42
21
7
386
2
0.4
10
Cascades
146
5.95
7.02
7.01
4
177
109
4
117
-
17
10
115
1
0.1
7
Idaho Batholith
68
6.32
7.19
7.01
19
288
103
12
204
64
12
9
43
1
0.3
7
NW Wyoming'
38
5.71
7.37
7.23
38
299
97
32
172
-
93
24
2909
4
0.4
32
Colorado Rockies
121
6.02
7.61
7.56
19
560
234
30
515
803
140
31
2212
2
0.5
17
Norway
202
4.48
5.60
-
-33
32
-
5
42
-
39
-
125
4
-
21
SkreSdalen
21
4.51
4.91
4.84
-31
-12
-14
7
13
13
41
42
52
10
10
10
Italian Alps
318
4.65
7.33
7.26
-22
460
186
15
453 215
157
64
5200
17
15
74
Ossola Valley
41
4.65
6 47
6.29
-22
193
31
15
315
81
160
58
1975
21
19
53
* Data from Landers et al. (1987)
6 Data from Wathne et al. (1900)
c excluding Fern Lake (4D3-017). which was affected by geothermal inputs
" calculated from pH
01
3
a
2
«T
0
(O
«
3
33

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(expressed as N03"), the base cations Ca2+ and Mg2+ were included in the CB term (Equation 1),
and assuming background N03' to be zero. They concluded that any strategy of pollutant
abatement in these areas would need to include nitrogen in addition to sulfur compounds.
However, Wathne et al. (1989) noted that the development of pollutant loads for the Alps related
to protection of lakes was difficult because of the large range in precipitation volumes
(80-200 cm). In this case, pollutant concentrations would be easier to address than would a
deposition-based standard.
Episodic acidification is an important issue for surface waters throughout high-elevation areas
of the West. A number of factors pre-dispose western systems to potential episodic effects,
including:
1.	the abundance of dilute to ultradilute lakes (i.e., those having extremely low
concentrations of dissolved solutes), exhibiting very low concentrations of base cations,
and therefore ANC, throughout the year;
2.	large snowpack accumulations at the high elevation sites, thus causing substantial
episodic acidification via the natural process of base cation dilution; and
3.	short retention times for many of the high-elevation drainage lakes, thus enabling
snowmelt to rapidly flush lake basins with highly dilute meltwater.
Thus, the physical characteristics (e.g., bedrock geology, lake morphometry) and climate
throughout high elevation areas of the West provide justification for considering potential episodic
acidification to be an important concern. In addition, the few studies that have been conducted to
date confirm the general sensitivity of western lakes to episodic processes.
In the Rocky Mountain region, Loch Vale has been the subject of intensive research on
hydrochemical responses to snowmelt. Loch Vale is located in Rocky Mountain National Park,
CO, at an elevation of 3100 to 3900 m. The watershed is comprised primarily (81%) of exposed
bedrock and talus (Baron and Bricker 1987). The snowmelt is characterized by three general
periods:

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1.	flushing of concentrated lakewater out of the lake during early snowmelt, and release of
the winter's accumulated atmospheric deposition from the snowpack. The minimum
lakewater pH (~ 5.7) and highest N03 concentrations occur at this time.
2.	dilution phase during which ionic concentrations in lakewater are diluted by the
continuing melt,
3.	decreasing hydrologic flows, yielding very dilute lakewater during the summer months
(Baron and Bricker 1987, Wigington et al. 1990).
High concentrations of N03" and S042" in the early phases of snowmelt have also been
documented for other high-elevation sites in Colorado (Reddy and Caine 1988) and Wyoming
(Clow et al. 1988).
Baron and Denning (in press) examined seasonal patterns of wet deposition chemistry at
Loch Vale, and also a nearby lower-elevation (2500 m) site, Beaver Meadow. Zonal wind and
solute association data were interpreted by Baron and Denning (in press) as evidence that strong
acid anions, NH/ and high salt concentrations originated primarily to the east of Rocky Mountain
National Park and were transported via upslope winds or convective instability from differential
heating at the higher and lower elevations. These influenced the composition of precipitation at
the high elevation site most strongly during summer. Winter precipitation was derived primarily in
conjunction with westerly zonal winds, resulting in orographic or cyclonic precipitation (Baron and
Denning, in press). Potential sources of strong acid anions to the west of the park were not
investigated, although winter precipitation was generally more dilute.
Vertucci and Corn (in press) collected weekly water samples during the snowmelt season at
three sites in the Front Range, Colorado, ranging in elevation from 2800 to 3050 m. They also
examined data from sites in the Elk Mountains, Colorado, and West Glacier Lake, Wyoming.
Snowmelt adjacent to West Glacier Lake was acidic (ANC = -20 //eq L'*), although water sampled
from the lake outlet maintained ANC > 0 on nearly all sampling occasions. Patterns of ANC at
the Elk Mountain site were related to changes in base cation concentrations (dilution). Vertucci
and Corn (in press) found no evidence of acidic conditions at any of the'three Front Range sites,
March, 1994
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and minimum ANC values were associated with low base cation concentrations, rather than
elevated acid anion concentrations.
In the Elk Mountains of central Colorado, Harte and Hoffman (1989) hypothesized that
episodic acidification was responsible for the decline of tiger salamanders (Ambystoma tigrinum)
in high elevation ponds. Based on a review of these data (Blanchard et al. 1987, Wissinger and
Whiteman 1992) and snowmelt chemistry from other Colorado sites, Vertucci and Corn (In Press)
concluded that current deposition chemistry in the study area was insufficiently acidic to have
caused runoff of strong acid anions in concentrations great enough to cause mortality of
amphibians, particularly considering that the snowmelt preceded the onset of amphibian
reproduction by at least several weeks. In the Sierra Nevada, no relationship was found linking
current deposition chemistry and the status of amphibian populations (Bradford et al. 1992,
Soiseth 1992).
Two high altitude (3200-3700 m) catchments in southern Wyoming were monitored for major
ions from 1988 through 1990 (Reuss et al. In Press). The volume-weighted pH of deposition at
the study sites was pH 5.4 and N03' and S042' concentrations in the snow averaged about 10
fieq L'\ The ANC in the two lakes, East Glacier Lake and West Glacier Lake, averaged 50 and 39
fieq L'\ respectively, and decreased by 10 to 20 fieq L'1 during snowmelt. The two inlets to West
Glacier Lake averaged ANC less than 10 /*eq L"\ yet maintained a positive ANC throughout the
study period. The substantially higher ANC values in the lakes relative to measured inlet stream
values were attributed to unmeasured groundwater inputs. In East Glacier Lake, which lacked
permanent inlets, nitrogen uptake was highly efficient, resulting in removal of 96% of the nitrate
input to the watershed. In contrast, only 60% of the nitrate was removed from the West Glacier
Lake watershed. Using the empirical Henriksen model,
A ANC = A S042" - ACB
= A S042" - (F * A SO,2)
(2)

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where F = Sin 90 (CB*/S) and S is the ANC at which F = 1.0 and is expected to be from 200 to
400 peq L'\ Reuss et al. (in press) calculated that the precipitation pH would need to decrease to
within the range of 4.11 to 4.33 in order to chronically acidify East Glacier Lake and to a pH of
4.20 to 4.35 to acidify West Glacier Lake. Obviously, episodic acidification would be expected to
be an important consideration long before these lakes would become chronically acidic. To
acidify the streams would require an estimated increase of <10 fieq L'1 of acid, which
corresponds to a precipitation pH of 4.9. Thus, within these two adjacent catchments, we
observe a considerable range in expected response of aquatic systems between the two study
lakes, and in particular between lakes and streams.
In the North Cascades, Loranger and Brakke (1988) observed a 50% decline in the ANC of
Bagley Lake to a low of 67 ,ueq L"1 following snowmelt. Because the anion concentrations were
so low in the snowpack, most of the ANC loss was attributed to dilution of base cations by
meltwater. The annual average ANC of Bagley Lake was 100 ^eq L'\ From August through
October, the ANC in the lake increased by nearly 30 fieq LThe shallow depth of the lake
promoted rapid mixing and flushing (short residence time). Thus, most of the nitrate in the
snowpack passed through Bagley Lake (elev. 1800 m) during the period of most rapid snowmelt;
by mid-June nitrate concentrations in the lake had decreased from over 5 /zeq L'1 to less than 2
/xeq L'1. Loranger and Brakke (1988) concluded that NOa" and S042' concentrations in the
snowpack were too low to result in significant pH depression during snowmelt. In contrast, Eilers
and Bernert (1994) coserved virtually no change in ANC in a small lake in the southern Cascades
during snowmelt. Lake Notasha, a seepage lake in southern Oregon also situated at 1800 m,
varied only a few microequivalents from an average ANC of 10 ueq L"1. Despite its extremely low
CB (~ 20 jieq L'1; Eilers et al. 1990), the lack of surface inlets to Lake Notasha results in a long
hydraulic residence time (5-10 yrs) which provides a buffer against even substantial increases in
acidic inputs. Because of high anion retention, the lakes in the southern Cascades are likely far

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less sensitive to potential increases in acidic deposition than lakes with similar CB elsewhere in
the West.
Welch et al. (1992) observed a statistically significant decrease in lakewater sulfate during
1983 through 1988 in six high elevation Cascade Mountain lakes. The total decrease in sulfate
concentration averaged 2.2 ^eq L"' in slow-flushing lakes and 4.2 /*eq L' in fast-flushing lakes.
Coincident with these changes in lakewater sulfate concentrations was a sharp decrease in S02
emissions from the ASARCO smelter, located 100 km SW of the lakes, from 87 kt yr"1 in 1983 to
12 kt yr'1 in 1985 (the year of its closure), and 0 thereafter. Emissions of S02 from Mt. St. Helens
(~ 160 km to the south) also decreased by about 34 kt yr'1 from 1983 to 1988. The fastest
decreases in lakewater nonmarine sulfate concentrations occurred from 1983 to 1984 in fast-
flushing lakes (~ 2 ueq L"1) and from 1985 to 1986 in slow-flushing lakes (~ 1 /zeq L"'). This
observation, coupled with the larger change in emissions from the smelter (as compared with Mt.
St. Helens) and location of the smelter relative to prevailing wind directions, suggest that the
ASARCO closure was the principal cause of the observed decline in lakewater S042"
concentration. Lake alkalinity did not show a consistent change in response to the observed
decrease in lakewater sulfate concentration (Welch et al. 1992).
In contrast to lakes in the southern Cascades, lakes in the Sierra Nevada receive both greater
acidic deposition and higher rates of runoff. In the Emerald Lake watershed (elev. = 2800 to
3416 m), N03" increases of about 120% have been observed in the streams during snowmelt
(Melack and Stoddard 1991, Williams and Melack 1991b). Concentrations of N03 in the Emerald
Lake outlet increased from 2-3 fieq L'1 in the fall to 10-13 //eq L'1 during spring runoff. The
increase in N03 and S042" (~ 50%) was attributed to preferential elutriation from the snowpack
and low retention rates in the watershed. Reduction of N03" and S042" within Emerald Lake was
relatively small, and most of the acid anions passed through the lake outlet. This increase of both
S042' and N03' in the snowmelt in the Sierra Nevada contrasts with observations in the

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Adirondacks where snowmelt runoff diluted SO/ as well as base cation concentrations (Schaefer
etal. 1990).
Melack et al. (1993) reported two years of intensive research at seven high-elevation lakes in
the Sierra Nevada. The lake elevations varied from 2475 m to 3425 m, and the catchments
spanned the length of the high-elevation Sierra Nevada. Solute concentrations, particularly ANC
and CB, were greatest during winter, declined to minima during snowmelt, and gradually
increased during summer and autumn. Sulfate concentrations varied most in lakes with lowest
volumes, and were generally less than 10 peq L ' in five of the seven lakes. In Speuller Lake and
Topaz Lake, however, S042' concentrations generally ranged from 10 to 30 ^
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minerals, or changing dilution of S042' by snowmelt. Based on lake concentrations of S04z" and
CI' and on wet deposition concentrations of S042', N03", and H\ Turk and Spahr (1991)
concluded that low ANC lakes have lost no more than 5 peq L"' ANC in the Bitterroot Range of
the Northern Rocky Mountains, 12 /zeq L1 ANC in the Wind River Range of Wyoming, and 10 fieq
L"' ANC in the Front Range of Colorado.
Summit Lake located at 1650 m in the Western Washington Cascades was originally sampled
by Landers et al. (1987) and in 1993 by the Mt. Baker-Snoqualmie National Forest (Eilers and
Bernert, unpublished data). The current lake ANC is 1 ^eq L"1 with a nonmarine [S042] of 7 fieq
L'\ Because the nonmarine [CB] is only 10 ^eq L'1 and can largely be attributed to deposition,
there appears to be little opportunity for watershed contribution of S. Summit Lake is close to
local and regional sources of S emissions and it is conceivable that the lake has lost up to 7 to 10
fieq L'1 of ANC in response to atmospheric deposition.
b. Paleolimnoloav
Diatom-inferred pH and ANC were calculated at 24 depth intervals at Emerald Lake,
California, for the period 1825 to the present (Holmes et al. 1989). Emerald Lake is a very dilute,
high-elevation lake in Sequoia National Park. Significant trends were not found for either pH or
ANC, and the authors concluded that Emerald Lake has not been affected by acidic deposition.
Whiting et al. (1989) completed paleolimnological analyses of three additional lakes in the Sierra
Nevada. Eastern Book Lake (pH = 7.06) showed evidence of both long-term alkalization (~ 0.3
pH units over the past 200 years) and pH fluctuations since 1970. Lake 45 (pH = 5.16) may have
acidified slightly (~ 0.2 pH units) over the last 60 years. Lake Harriet (pH = 6.52) showed no
significant change. Baron et al. (1986) investigated metal stratigraphy, diatom stratigraphy, and
inferred pH profiles of four subalpine lakes in Rocky Mountain National Park, Colorado. The
authors of this study also found no evidence of historical influence on pH attributable to
atmospheric deposition. Eilers and Dixit (1992) reported a slight dilution or acidification trend in

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Lake Notasha, OR (ANC = 10 peq L'1), based on recent changes in diatom stratigraphy.
However, there was no diatom calibration set available to allow quantitative estimates of change
in lake chemistry. Thus, the limited paleolimnological data available for lakes in the western
United States suggest that widespread chronic acidification probably has not occurred.
c. Process models
The Forest Service, as part of its program to evaluate air quality related values in wilderness
areas, initiated a project to assess the sensitivity of wilderness lakes to acidic deposition.
The sensitivity of selected lakes in Idaho and Montana was assessed by modeling lake response
to simulated increases in sulfate deposition (Eilers et al. 1991). Fifteen lakes with alkalinity values
ranging from 21 to 835 ,ueq L'1 were selected for study from three wilderness areas, the Selway-
Bitterroot (SBW), Absaroka-Beartooth (ABW), and the Bob Marshall (BMW). The MAGIC model
was calibrated to each watershed using area-wide estimates of deposition and soil characteristics.
Current wet deposition of sulfur is approximately 2 kg S ha'1 yr'1.
The projected chemistry of the study lakes using MAGIC suggested that the lowest alkalinity
lakes would remain non-acidic (alkalinity > 0) if the sulfate deposition load were tripled over
present values, e.g., to about 6 kg S ha"' yr"1. At a five-fold increase in sulfate deposition
(10 kg S ha'1 yr'1) two of the six lakes modeled in the SBW were projected to become chronically
acidic. None of the lakes in the ABW or BMW were projected to become acidic at this five-fold
loading, although one lake in the ABW exhibited a projected alkalinity of 4 peq L'1 and a pH of
5.65. The lakes in the SBW are largely subalpine systems and generally contain moderate soil
cover. Nevertheless, the relatively unreactive granitic parent material, coupled with high rates of
precipitation, make these lakes among the most likely in the Northern Rocky Mountains to
become acidic under increased deposition loads. Although the watersheds in the ABW had a
higher percentage of exposed bedrock than watersheds in the SBW and BMW and showed the
largest decreases in modeled alkalinity, the weathering of rocks in the ABW is apparently

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sufficient to maintain positive alkalinities under moderate to heavy acidic deposition loads. The
lakes modeled in the BMW are located on limestone and argilitic sedimentary bedrock. These
watersheds have sufficient base cation production to ensure that the lakes will not acidify under
any realistic increase in acidic deposition.
The MAGIC modeling results provide estimates of long-term lake response to sustained
increases in sulfur deposition. It is likely, however, that low alkalinity lakes in the western United
States will acidify under episodic conditions before chronic acidification is evident. Estimates of
both episodic and chronic acidification potential could be refined by collecting more watershed-
specific information on low alkalinity lakes in the SBW and ABW. More detailed information is
needed on soil properties, hydrologic flowpaths, weathering characteristics of bedrock,
deposition, and surface water chemistry during snowmelt.
Eilers et al. (1991) reported the F-factors, or ratio of increase in base cation concentration
compared to the increase in acid anion (S042' + N03) concentration, calculated from MAGIC
model projections, using a three-fold increase in current S042 deposition. The reported F-factors
for individual lakes ranged from near zero to 0.65. Subregional average F-factors ranged from
0.25	in the Absaroka-Beartooth Wilderness to 0.43 in the Selway-Bitterroot Wilderness (Eilers et al.
1991).
B. UPPER MIDWEST
1.	Characteristics and Sensitivity to Acid Deposition
The Upper Midwest is characterized by numerous lakes created by repeated glaciations. The
region shows little topographic relief and the deep glacial overburden results in little or no
exposed bedrock. Sensitive aquatic resources in the Upper Midwest are largely comprised of
seepage lakes (Eilers et al. 1983). Limited studies of stream systems have shown them to be
insensitive to acidic deposition because groundwater discharge constitutes a major component of
stream flow. Furthermore, because of the low relief and permeable till, streams in this region are

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not subject to abnormally high discharge in the spring such as observed in the West (Eilers and
Bernert 1989). The seasonal depression of pH observed in some Wisconsin lakes cannot be
attributed to snowmelt runoff, but rather is caused by C02 accumulation from respiratory
processes under the ice (Kratz et al. 1987). Most drainage lakes (and some seepage lakes) in
the Midwest receive much of their inflow from groundwater which is generally enriched in ANC
from dissolution of carbonates and silicate minerals. Those seepage lakes with low base cation
concentrations receive nearly all of their hydrologic inputs as precipitation directly on the lake
surface (Baker et al. 1991a). Consequently, these lakes generally have long hydraulic residence
times (e.g. rw = 10 yr). The long tw provides an opportunity for in-lake reduction and assimilation
processes to neutralize much of the acidic inputs which would otherwise be concentrated from
evaporation. Baker and Brezonik (1988) quantified the importance of in-lake processes for sulfate
reduction as a function of rw in their IAG (Internal Alkalinity Generation) Model. For those lakes
that receive the vast majority of their hydrologic input from on-lake precipitation, even small
additions of groundwater can be expected to neutralize acidic deposition from atmospheric
sources indefinitely (Kenoyer and Anderson 1989). Only where groundwater contributions
approach zero can one expect that acidification from atmospheric sources would be a serious
concern in the Upper Midwest.
The critical importance of hydrologic flowpaths in the Upper Midwest has been refined and
reiterated in a number of regional assessments (Eilers et al. 1983, Nichols and McRoberts 1986,
Linthurst et al. 1986), synthesis analyses of this region (Baker et al. 1990a, Cook and Jager 1991),
and process studies of low-ANC lake systems (Lin and Schnoor 1986; Kenoyer 1986; Anderson
and Bowser 1986; Kenoyer and Anderson 1989, Wentz and Rose 1989; Webster et al. 1987, 1990,
1993; Webster and Eilers* in press).
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2. Current and Projected Future Deposition
Emissions of S02 and NOx in the Upper Midwest appear to have increased dramatically in the
20th century, reaching a maximum in the 1970's. Recent trends indicate that emissions reached
a maximum in about 1978 and have decreased since that period (Placet et al. 1990). In
Wisconsin and Minnesota, sulfur emissions decreased in the mid-1980's, following enactment of
pollution control legislation. However, the greatest reduction in regional emissions is attributed to
a reduction of S02 emissions in Michigan, which declined from about 1.3 M tons/yr in 1975 to
about 0.6 M tons/yr in 1988 (Webster et al. 1993). Recent declines in NOx emissions appear to
be substantially less than those observed for S02. We expect that S02 emissions will continue to
decline as older stationary sources are retired; however, NOx emissions appear to have stabilized
and can be expected to follow trends related to future consumption of fossil fuels within the
region.
Wet sulfur deposition ranges across the region from about 3 to 4 kg S ha"' yr' in Minnesota
to near 5 kg S ha' yr' in eastern portions of the Upper Peninsula of Michigan (Figure IV.A.1).
Nitrogen deposition follows a similar pattern and ranges across the region from about 3 to 4 kg N
ha' yr"' in wet deposition. These levels of deposition are moderate compared to other regions of
the country.
Deposition of base cations is an important component influencing the acidity of deposition in
the Upper Midwest and the major ion chemistry of seepage lakes. One of the major sources of
atmospheric base cations is dust derived from unpaved roads (Placet et al. 1990). As more rural
roads continue to be paved, it is likely that atmospheric concentrations of base cations will
continue to decline, thus partially negating the positive effects of expected continued declines in
S02 emissions.

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3. Current Surface Water Chemistry
Lakes in the Upper Midwest exhibit a considerable spatial pattern; lake population estimates
tor pH, ANC, base cations, and DOC decrease from west to east (Table IV.A.4). Sulfate
concentrations in lakes, however, do not show a comparable change across the region when all
lakes are included in the analysis, despite a 40% increase in wet S04 deposition from Wisconsin
to Michigan. Cook and Jager (1991) attributed the major west-to-east pattern in lake chemistry in
the Upper Midwest to increasing frequency of seepage lakes in the eastern portion of the region,
rather than a gradient in the deposition of sulfur. They attributed the absence of a more
pronounced gradient in lakewater sulfate concentration across the region to a combination of
processes, including watershed sources of sulfur (primarily in Minnesota) and variable anion
retention in lakes, related to supply of iron and organic carbon. The retention of sulfate by
dissimilatory reduction is particularly high for seepage lakes. For example, a seepage lake with
mean depth of 3 m and rw of 7.5 years would be expected to lose about 50 fieq L"' of S042' from
the water column by this process (Cook and Jager 1991). In contrast, a typical drainage lake
(2.5 m deep, rw = 1 yr) in the Upper Midwest would be expected to lose only about 10 ^eq L"1 of
SO,2"
Concentrations of inorganic nitrogen are uniformly low throughout the Upper Midwest and
are efficiently retained in both terrestrial and aquatic systems. Snowmelt does not provide any
significant nitrate influx to lakes in the Upper Midwest because most snowmelt first infiltrates into
the soil prior to reaching the drainage lakes. Snowmelt inputs of N into seepage lakes is limited
mainly to the snow on the lake surface. Aluminum concentrations are far lower in the Upper
Midwest than observed in lakes of similar pH in the Northeast, at least in part because of the
absence of significant watershed inputs of N to seepage lakes.
The large number and extent of wetlands in the Upper Midwest contribute to high production
of organic matter which is reflected in high DOC concentrations in many surface waters.
Nevertheless, base cation production is the dominant ion-enrichment process in most lakes in the

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Table IV.A.4. Population estimates ol selected parameters for lakes in the Upper Midwest. Units for ANC, CB, and S042 are in
microequivalents per liter. DOC is in milligrams per liter.
pH	ANC	CB	SO/	DOC
Population	n	N	ps p^	p5 Pjq	p5 p^,	P 50 P#s	P^ P8S
Northwestern	117 1519 6.11 6.68 31 126 113 232	45	76	9.2 15.3
Wisconsin (2c)
Northcentral	218 2368 5.31 7.03 -1 278 80 429	62 223	5.1 13.8
Wisconsin (2d)
Upper Peninsula	62	540 4.53 7.41 -34 904	55 1028	83 246	6.8 20.1
Of Michigan (2e)

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Upper Midwest. Even in low-ANC groundwater-recharge seepage lakes, base cation production
accounts for 72% to 86% of total ANC production (Cook and Jager 1991). Sulfate is the dominant
anion in the low-ANC (< 50 ^eq L'1) groundwater recharge lakes.
Episodic processes have been studied in two seepage lakes, Lakes Clara and Vandercook,
in northern Wisconsin (Schnoor et al. 1986; Webster and Eilers, In Press). Lakewater ANC
decreased in both lakes by about 6 fxeq L"1 in response to a large summer rainstorm. Kratz et al.
(1987) reported spring episodes at Little Rock Lake, Wl. They concluded that most meltwater
flows into the till surrounding this seepage lake, and consequently does not directly impact the
lake to any degree. Springtime pH depressions were attributed primarily to increased P^
generation under ice cover, in response to hypolimnetic decomposition and respiration. In Upper
Michigan, McNearney Lake was the subject of monthly sampling by Cadle et al. (1984), who
reported increased concentrations of S042', H+, N03, and NH4+ during snowmelt. Base cations
decreased slightly.
The Upper Midwest is a diverse region, with pronounced spatial gradients from west to east
in hydrologic lake type, major ion chemistry, and sulfur deposition. Lakes within the easternmost
portion of the region receive the highest sulfur deposition and also exhibit the greatest sensitivity
to potential acidification. Although deposition in the Upper Midwest has been declining in recent
years, those lakes with ANC near zero could be expected to acidify if trends in S emissions
reversed.
4. Quantitative Assessment of Acidification
The Upper Midwest has a large population of low ANC lakes, but relatively few acidic
(ANC < 0) lakes (Linthurst et al. 1986). Paleolimnological evidence suggests slight
acidification of selected lakes (Garrison 1990, Kingston et al. 1990) consistent with the modest
historical and current rates of sulfur deposition. Time-trend analysis of 28 lakes in the region
showed decreasing S042' concentrations in the lakes from 1983-1989, consistent with decreases

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in regional S02 emissions and sulfur deposition during the period (Webster et al. 1993). The
most apparent trends in lake chemistry were observed in the higher ANC lakes, which were
generally short-residence time drainage lakes. Most lakes failed to show significant trends in
acid-base chemistry other than S042 concentration, perhaps because:
1.	the period of the time-series was too short to allow for the chemistry to reach equilibrium
with changing atmospheric deposition in the long-residence time lakes, or
2.	changes in ANC were minimized by other changes in the acid-base chemistry of the
lakes (e.g., organics, aluminum, or base cations) that were too small to detect.
a. Space-for-time substitution
Most of the acidic and low ANC lakes in the Upper Midwest are seepage type. Some
drainage lakes with ANC <. 50 jieq L"' were sampled in ELS-I, however, and they exhibited a
pattern of increasing [S042;]/[CB] and decreasing [HC03' - H+] across a longitudinal gradient from
eastern Minnesota to eastern Michigan (Figure IV.B.1, Sullivan 1990). Atmospheric deposition of
sulfur and hydrogen ion increase along the same gradient (Eilers et al. 1988a). The decline in
[HC03 - H+] cannot be explained by differences in concentration of DOC, but is driven by four
acidic lakes in eastern Michigan that have very low [Ca2+ + Mg2+] (Figure IV.B.1). Three of the
four lakes had [S042']/[CB] between 0.7 and 0.9 and only moderate concentrations of DOC (~ 3
to 6 mg L'). High [S042] relative to [CB] is largely responsible for the current acidity of these
systems. The fourth lake has [S042]/[CB] of only about 0.5 and much higher DOC (~" 13 mg L ').
Organic acidity plays a major role in the current acidity of this lake and also of the lowest ANC
lake sampled in ELS-I in central Wisconsin (DOC ~ 16 mg L. ') (Figure IV.B.1). These spatial
distributions for low ANC drainage lakes in the Upper Midwest are consistent with the hypothesis
of historical acidification of some lakes with very low [Ca2+ + Mg2+] in the easternmost portion of
the region. For some of the lowest ANC systems, however, organic acidity is also an important
factor (Sullivan 1990).

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40
E. Minnesota	E. Michigan
Central Wisconsin
o
o
C2.
10
CVI
CM
O
o
100 H
Q
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The distribution of selected chemical concentrations across a longitudinal gradient in the
Upper Midwest is presented in Figure IV.B.2 for low ANC (< 50 jieq L') groundwater
recharge seepage lakes (Sullivan 1990). This hydrologic lake type was defined, on the basis of
low (< 1 mg L') silica, which is indicative of minimal weathering (thus groundwater) inputs.
Lakes having high [CI ] (> 20 p.eq L') were deleted to minimize road salt influence and other
anthropogenic disturbances. Although biological uptake of Si02, for example by diatoms,
influences surface water concentrations, [Si02] in seepage lakes is generally indicative of
groundwater input (Sullivan 1990, Baker et al. 1991a).
Across an increasing sulfur depositional gradient from eastern Minnesota eastward to eastern
Michigan, [HC03" - H+] decreases and the ratio [S042 ] to [CB] increases in these groundwater
recharge seepage lakes (Figure IV.B.2). In Michigan and Wisconsin, many lakes have [S042] >
[C0] and are currently acidic because of high [S042] relative to [CB]. There are also many lakes
that have high concentrations of DOC, and organic acidity undoubtedly accounts for many of
these lakes having ANC <_ 0, particularly where [CB] is low (e.g., < 50 fieq L"'). The spatial
pattern in [HC03 - H+] is not attributable to DOC, which generally shows a decreasing trend with
increasing deposition. [Ca2+ + Mg2+] also decreases with increasing deposition, and this is
probably attributable to lower levels of base cation deposition and greater amounts of
precipitation in the eastern portion of the region. Atmospheric deposition is an important source
of base cations for groundwater recharge seepage lakes because of minimal groundwater inputs.
In the eastern portion of the region, these lakes are more sensitive to pH and ANC depression in
response to either elevated [SO„2] or DOC. The spatial pattern for low ANC groundwater
recharge lakes in the Upper Midwest are consistent with the following hypotheses (Sullivan 1990):
1. Sensitivity to mineral and organic acidity as mechanisms for depressing ANC below zero
increases from west to east because of decreasing lakewater [CB], and this may be due,
in part, to changes in base cation deposition and precipitation volume along this
gradient.

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50
E. Minnesota	E. Michigan
Cenfral Wisconsin
i
I ro
O
O
0 -
-50 -4
m
o
1.0-
i

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2.	High concentrations of DOC are responsible for the acidic conditions in some of the
lakes, and DOC may have decreased in response to acidic deposition.
3.	Many of the lakes in eastern Michigan, and some in Wisconsin, are currently acidic
because of high [SO,2] relative to [CB], and have probably been acidified by
atmospheric deposition.
There are few data available with which to evaluate the acid-base chemistry of seepage lakes
in the absence of acidic deposition. Although sulfur deposition is low in the western portions of
the Upper Midwest region (i.e. Minnesota), the predominant hydrologic lake type in this part of
the region is drainage. Seepage lakes are more common in Wisconsin and Michigan, where
sulfur deposition is considerably higher.
Perhaps the best data set available for evaluation of relatively pristine seepage lake chemistry
and comparison with Upper Midwestern seepage lakes that receive higher deposition was
collected on the Kenai Peninsula in Alaska. A probability sample of lakes on the Kenai Peninsula
was conducted in August 1988, when 59 randomly selected lakes were sampled from an
estimated population size of over 800 lakes (Eilers et al. 1993). The study area was the glaciated
plain on the northwestern portion of the Kenai Peninsula. Kenai lakes were similar to lakes in the
Upper Midwest in that both were in areas of glacial till and outwash and both areas have a high
percentage of seepage lakes. Forty-seven of the 59 lakes sampled in the Kenai were seepage
systems and with few exceptions the seepage lakes had low concentrations of ANC, CB, and
silica. Although lakes in both regions had similar distributions of conductivity and color, only
6.8% of the lakes in the Kenai had pH <_ 6.0 and none had pH <. 5.0, compared to 27.7% and
2.1%, respectively, in the Upper Midwest. None of the lakes sampled in the Kenai Peninsula were
acidic (ANC 0). Dissolved organic carbon (DOC) concentrations in the lakes ranged from 1.3
to 17.4 mg L'\ with a median concentration of 5.5 mg L \ The maximum S042" concentration
measured was 45 ueq L1 with a median of 3 ueq L'1; the median sea salt corrected sulfate
concentration was about 0 ueq L'. The lakes with higher sulfate concentrations (i.e.,

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> 10 ueq L"') generally had high ANC values (e.g., > 1 meq/L). Sea salt corrected base cation
concentrations in the lakes ranged from 48 to 1945 fieq L', with a median of 200 ^eq L'\ Plots of
ANC + organic anion (as estimated from anion deficit) versus C'a showed that the lakes fit the
1:1 line as expected for dilute bicarbonate lakes unacidified from atmospheric deposition (Eilers et
al. 1993). The Kenai lake data set is important because it is one of the few data sets of seepage
lake chemistry in an area receiving low levels of acidic deposition in the western hemisphere.
The southern Cascades also contains a group of seepage lakes that receive low levels of acidic
deposition (e.g. Eilers et al. 1990).
b. Monitoring
Eilers et al. (1989) selected a set of data from 145 of the northern Wisconsin lakes that had
been surveyed by Birge, Juday, and coworkers during the period 1925 to 1941. After correction
of historical data to enable direct comparison with recent measurements by Eilers et al. (1983),
the data suggested, in general, that conductivity, [CaJ+], pH, and alkalinity had increased during
the intervening 50-year period. The increases were attributed to watershed development, and in
two cases to major watershed fires. When only the 19 lakes that had not experienced
development were considered, it was estimated that 7 had experienced a pH decline. Because of
the uncertainties involved, the authors could not attribute the apparent decline in pH of some
iakes to a specific cause.
Webster et al. (1993) presented data on temporal trends in the chemistry of 28 Upper
Midwestern lakes sampled three times per year between 1983 and 1989. The lakes are located in
Minnesota (4), Wisconsin (13), and Michigan (11). Statistically-significant trends in S042'
concentration were found for eight lakes and were negative in direction, consistent with a recent
regional decline in sulfur emissions and deposition. Lakes exhibiting significant declines in S042'
concentration were primarily drainage type and had ANC >100 ueq L"\ only one, McNearney

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Lake, was seepage type and low in ANC (-39 fieq L''). Furthermore, of the eight lakes that
showed decreases in S042 concentration, only Buckeye Lake also showed increases in ANC
(+2.4 fieq ANC L'1 yr'), and this lake was relatively high in pH (~ 7.1) and ANC (168 fieq L"').
For the most part, the anticipated trends of decreasing S042' and increasing pH and ANC that
were expected to occur in the more sensitive acidic and low-ANC Upper Midwestern seepage
lakes were not found. Although 15 seepage lakes having pH < 6 were included in the monitoring
program in Wisconsin and Michigan, none of them showed the expected patterns of "recovery"
from acidic deposition effects. During the study period, S042' deposition decreased about 9 to 14
eq ha'1 yr"' at three NADP monitoring stations in the region (about 4% per year decrease)
(Webster et al. 1993).
A second group of five lakes, primarily acidic or low-ANC seepage or closed-basin lakes,
exhibited significant trends of decreasing ANC, which were associated with concurrent declines in
base cation concentrations, rather than changes in S042' concentration. Nevins Lake, Michigan,
exhibited the most marked changes; ANC declined from 178 fieq L'1 in 1983 to 21 peq L"1 in 1988
and pH declined 0.75 pH units (Webster et al. 1990, 1993). These changes were attributed by the
authors to climatic fluctuations. Lower than normal precipitation volumes reduced lake and
groundwater levels, resulting in diminished inflows of ANC-rich groundwater into these (primarily
seepage) lakes. The magnitude of the ANC decrease exceeded the magnitude of lake
acidification from atmospheric deposition by a considerable margin. Thus, if the acidic and low-
ANC seepage lakes were responding to decreases in acidic deposition, such responses were
overshadowed by changes in acid-base chemistry related to climate and hydrology.
c. Paleolimnoloqy
In the Upper Midwest region, diatom-inferred pH reconstructions have been completed for 15
lakes, and summarized by Cook and Jager (1991). Four lakes, all of which have current pH
< 5.7, showed a pH decline of 0.2 to 0.5 pH units during the past 50 to 100 years.

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Diatom-inferred pH increased in one lake by 0.2 pH units. No change was inferred for the other
10 lakes, including 4 lakes with current pH > 6.0. Nine of the lakes were included in PIRLA-I, and
results from these have been described by Kingston et al. (1990). No major, recent, regional
acidification was indicated by the diatom-inferred pH reconstructions. Changes in most lakes
were apparently small, and were no greater during the industrial period than during the pre-
industrial period. Two lakes were studied in Michigan in PIRLA-I, McNearney Lake and Andrus
Lake. McNearney Lake is highly acidic (ANC = -38 ^eq L'1) and has experienced little change in
pH for thousands of years. It is atypical for upper midwestern lakes because of its extremely low
pH and ANC and high concentrations of labile monomeric Al. Its current pH is strongly buffered
by Al (Cook et al. 1990). The pH of Andrus Lake was inferred to have declined about 0.3 pH
units from 1840 to the turn of the century, followed by a rise of 0.2 units after major logging and
fire in the watershed around 1920. Subsequently, the pH declined about 0.2 pH units to the
present (Kingston et al. 1990). The floristic changes in Andrus Lake in recent decades were
consistent with lake acidification from atmospheric deposition and decline in DOC and/or
response to watershed disturbance.
Four lakes were studied by PIRLA-I in Wisconsin, all of which have current ANC near zero
and pH 5.2 to 5.7. Brown Lake showed the most consistent decline in diatom-inferred pH (~0.5
units) of any upper midwestern lake studied. None of the pH reconstructions for other Wisconsin
lakes suggested recent acidification except possibly the one for Denton Lake, but it fluctuated
widely. The diatom-inferred pH of Otto Mielke Lake increased 0.2 units in recent decades,
coincident with changes in fisheries management. Camp 12 Lake also exhibited evidence of
slight pH increase, shortly after major logging and slash fire in the watershed about 1912
(Kingston et al. 1990).
Paleolimnological evidence for recent lakewater acidification in the Upper Midwest as a result
of acidic deposition is strongest for Brown and Andrus Lakes. However, Kingston et al. (1990)
estimated that the change in alkalinity was only -6 to -8 ueq L'1 for these lakes. Such changes

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are small relative to natural variability (Schnoor et al. 1986; Kingston et al. 1990) and relative to
changes inferred for many Adirondack lakes (Sullivan et al. 1990a).
d.	Process models
Pollman and Sweets (1990) conducted IAG model hindcasts for seven Upper Midwestern
lakes. The magnitude of simulated historical acidification followed the gradient in sulfur
deposition across the region, with the largest declines in ANC predicted for lakes in the Upper
Peninsula of Michigan, -20 /zeq L'1 for Andrus Lake, and -37 peq L"' for McNearney Lake,
assuming F=1.0 in groundwater and a groundwater concentration factor of 1.25. In northern
Wisconsin, simulated historical acidification ranged from near zero (Denton Lake) to slight (Otto
Mielke Lake, -8 fieq L'; Brown Lake, -4 ,ueq L '; Camp 12 Lake, -4.ueq L '). Agreement between
IAG hindcast and diatom-inferred pH change was reasonable for five of the lakes. But for the
lakes with the largest simulated declines in pH (Andrus -0.96 and Otto Mielke -0.68 pH units), the
model estimates of historical acidification were substantially larger than the diatom inferences
(-0.4 and +0.4 pH units, respectively; Pollman and Sweets [1990]).
e.	Experimental Manipulation (Little Rock Lake)
Quantitative information on the acidification response of a seepage lake is available from the
Little Rock Lake Acidification Project. This whole-lake manipulation project was initiated in 1983,
and involved the controlled acidification of a two-basin low-ANC seepage lake in northcentral
Wisconsin. The lake was divided into treatment and reference basins using a flexible vinyl curtain
and the treatment basin was acidified with sulfuric acid in steps of 0.5 pH units every two years,
from an initial pH of 6.6 in 1985.
The study was described by Brezonik et al. (1986) and Watras and Frost (1989). Chemical
results of the manipulation were presented by Sampson et al. (1993) and Brezonik et al. (1993).
Acidification to target values of pH 5.1 and 4.7 was achieved after addition of sufficient sulfur to

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raise the lakewater S042' concentration in stages from 53 to 116 and 147 /ieq L'\ In response to
the increased SO„2' concentration, ANC decreased from 27 fieq L"' to -4 and -14 peq L'1,
respectively (Brezonik et al. 1993). In both cases, increased mobilization of base cations,
presumably from the lake sediments, neutralized 53% of the added S042' (Sampson et al. 1993).
It is important to note that this degree of base cation neutralization (F=0.53) does not include any
watershed neutralization that might have occurred if the acid had been applied beyond the
boundaries of the lake itself, although groundwater contributions to Little Rock Lake during the
study approached zero, providing some justification for use of an in-lake neutralization F factor.
Nevertheless, this estimate of base cation neutralization should be viewed as conservative for a
seepage lake.
C. NORTHERN FLORIDA
1. Characteristics and Sensitivity to Acid Deposition
Florida lakes are located in marine sands overlying carbonate bedrock. Underneath the
Pliocene to principally Pleistocene age sands is the Floridan aquifer, an extensive series of
limestone and dolomite that underlies virtually all of Florida. In the Panhandle and northcentral
lake districts, the Floridan aquifer is separated from the overlying sands by a confining layer
known as the Hawthorne formation. The major lake districts are located in karst terrain, and lakes
probably formed through dissolution of the underlying limestone followed by collapse or piping of"
surficial deposits into solution cavities (cf. Schmidt and Clark 1980; Arrington and Lindquist 1987).
Flow of water from the lakes is generally downward, recharging the Floridan aquifer. Variations in
lake stage differ from lake to lake in response to long-term trends in precipitation, and lakes with
direct hydraulic connections with the Floridan aquifer show considerably broader ranges in stage
compared to lakes where the connection is impaired (cf. Clark et al. 1964a; Hughes 1967). Base
cation enrichment appears to be small in most study lakes in Florida and ANC generation is due
primarily to in-lake anion reduction (S042' and NOj; Baker et al. 1988; Pollman and
March, 1994
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Canfield 1991). Retention of S042 by watershed soils may also be important. Where
groundwater interactions with the deeper aquifers are present, surface waters can be highly
alkaline. However, those lakes with hydrologic contributions from shallow aquifers in highly
weathered sands can be quite acidic and presumably sensitive to acidic deposition. As is the
case elsewhere, the key to understanding the potential response of Florida lakes to acid inputs is
related largely to knowledge of the hydrologic flow paths.
Topographic relief in Florida is minimal and attempts to relate groundwater contributing areas
to specific lakes is highly problematic (Pollman and Canfield 1991). Detailed studies, of low-ANC
lakes in northern Florida show that, unlike low-ANC lakes in the Upper Midwest, groundwater
contributions can represent the major hydrologic input. For example, Lake Five-0 in the
Panhandle receives the majority of its annual inflow from groundwater sources. An additional
anomaly with regard to the flow path is that water does not exit the lake through the opposing
shoreline, but rather passes vertically downward through the lake bottom. Despite the
considerable groundwater contributions to Lake Five-O, the pH (5.4), ANC (-4 fieq L'1), and
nonmarine base cation concentrations are low (Pollman et al. 1991). This reflects the highly
weathered nature, and low base saturation, of the sands through which the groundwater flows
before entering the lake.
Although evaporation plays a role in most regions in concentrating acidic inputs from
atmospheric deposition, the effect of evaporation is much greater in Florida than other low-ANC
regions of the United States. Annual pan evaporation measured at several stations ranged from
149 cm to 175 cm, increasing in a southerly direction. As a consequence, the.net precipitation in
the Panhandle is 50 to 100% greater than that in the Central Trail Ridge (Pollman and Canfield
1991).
In-lake processes are also important components influencing the chemistry of Florida lakes.
Baker et al. (1988) illustrated the importance of in-lake anion retention in generating ANC for
Florida lakes. Retention of inorganic N is nearly 100% and ANC generation from SO/' retention

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may approach 100 fieq L'1 in some Florida lakes (Pollman and Canfield 1991). Base cation
deposition and ammonium assimilation are the other primary influences on the acid-base status of
Clearwater lakes, other than atmospheric deposition and evapoconcentration.
2. Current and Projected Future Deposition
Current deposition in Florida is moderately acidic with volume-weighted mean (VWM) pH
ranging from 4.55-4.68 for the four northern FADS (Florida Acid Deposition Study) sites. Non-
marine sulfate VWM concentrations ranged from 19.8 to 22.9 ueq L"' and nitrate VWM
concentrations ranged from 9.5-11.1 ,ueq LAmmonium VWM concentrations ranged from 4.2 to
6.3 fieq L'\ Based on regional estimates of dry:wet ratios for Florida, dry deposition of S and N
are 70% and 96%, respectively, of wet values (Baker 1991). Total S and N depositon in parts of
the northcentral peninsula are therefore approximately 10 and 9 kg ha"' yr'1, respectively.
For the southeastern United States (defined as EPA Region 4), both sulfur and nitrogen
emissions showed only modest increases from 1900 to 1960 (Gschwandtner et al. 1985).
However, from 1960-1980, emissions increased approximately four-fold and then began
decreasing in the 1980's (Gschwandtner et al. 1985). More detailed analyses of recent trends in
regional emissions indicate that emissions peaked about 1978 and have declined slightly in the
following decade (Placet et al. 1990). Model analyses suggest that within-state sources
contributed about two-thirds of the total deposition of sulfur in 1983 (FCG ,1986), therefore it may
be more appropriate to examine the emissions for Florida alone. Emissions of S02 in Florida
fluctuated around 900,000 tonnes yr1 from 1976 through 1984, but have been projected to
increase by nearly 30% by the year 2000 (FDER 1984). If population increases continue in
Florida, it appears reasonable to assume that NO„ emissions also will continue to increase at a
similar pace.

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3. Current Surface Water Chemistry
Northern Florida contains the highest percentage of acidic lakes of any lake population in the
United States (Linthurst et al. 1986). Seventy-five percent of the Panhandle lakes are acidic, as
are 26% of the lakes in the northern peninsula (Table IV.C.1). This large population of acidic
lakes, combined with increasing emissions of S and N for the state, has stimulated numerous
investigations of the acid-base chemistry of these lakes (Pollman and Canfield 1991). Most of
these acidic lakes are Clearwater (DOC < 420 //mol) seepage lakes in which the dominant acid
anions are CI" and SO*2'. The most dilute group of lakes is found in the Panhandle, which
Pollman and Canfield (1991) attributed to higher precipitation, lower evaporation, and lower
watershed disturbance. The regional difference in evapoconcentration for Florida has two
opposing effects (Pollman and Canfield 1991). Concentrating an acidic solution increases its
acidity. However, increasing evaporation has an opposing effect on lake chemistry by affecting
lake hydrology. As evaporation increases, groundwater inflow also increases in importance and
provides a proportionally greater supply of base cations. Increasing evaporation also increases
the lake hydraulic residence time (r J, thus increasing the opportunity for dissimilatory S042'
reduction (Baker and Brezonik 1988). Nitrate and ammonium concentrations in lakes without
agricultural contributions (as estimated by IC < 15 ueq'L'1) are generally not measurable.
Retention of inorganic nitrogen is highly efficient in Florida lakes and contributing areas, similar to
lakes in the Upper Midwest.
Although concentrations of DOC are high in many Florida lakes, organic anions are generally
less important than S042' in the low-ANC and acidic lakes (Pollman and Canfield 1991).
Aluminum concentrations are very low in Florida lakes despite their high acidity. Although Aln+ is
mobilized in surficial soils (e.g., < 15 cm depth) by the acid loading from atmospheric deposition,
most of,the Af"* is removed from solution by precipitation and ion exchange reactions within 75
cm depths (Graetz et al. 1985), and relatively little Af+ is transported in solution to lakewaters.
March, 1994
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Table V.C.I. Population estimates of selected parameters for lake populations in Northern Florida." Units for ANC, CB, and S042 are in
microequivalents per liter. DOC is in milligrams per liter.



pH


ANC
cB
b

SO,2 b

DOC
Population
n'"
N
Ps
Pso
Ps
Pso
Ps
Pso
Pso
P«s
Pso
Pes
Panhandle
9
144
4.84
5.03
-8
-7
83
112
34
64
3 3
6.3
Northcentral
23
379
4.41
5.84
-23
0
194
284
94
339
1.9
13.6
" Source: Linthurst et al. 1986
0 Includes marine aerosols
>

o S
Si
ST ®
8 2
Be
(0
m
3
<
O
3 ~
a > a
3-g o
ES-g
0	° 2
=r 30 £
5*5'
1	g °
a c r-
^ s?
(A 2.
3 
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4. Quantitative Assessment of Acidification
Evidence for recent acidification of some Florida lakes has been supported by historical
analyses of lake chemistry (Crisman et al. 1980, Baker 1984, Brezonik et al. 1983, Battoe and
Lowe 1992), inferred historical deposition (Husar et al. 1991, Hendry and Brezonik 1984), and
paleolimnological reconstructions of lake pH (Sweets et al. 1990; Sweets 1992). However, the
case for acidification by acid deposition is equivocal with respect to all lines of evidence (cf.,
Pollman and Canfield 1991) and the interpretation is complicated by profound regional and local
changes in land use and hydrology. For example, an alternative explanation (other than acidic
deposition) for the apparent acidification of Lakes Barco and Suggs (Sweets et al. 1990) is that
the apparent recent decline in pH may have been caused by a regional decline in the
potentiometric surface of the groundwater. Large groundwater withdrawals of the Floridan aquifer
for residential and agricultural purposes may have contributed to reduced groundwater inflow of
base cations into seepage lakes, thereby causing lakewater acidification.
Other land use changes have probably increased lake pH by providing increased inputs of
fertilizer, thus increasing the productivity of many lakes. Paleolimnological evidence of this
process is provided by Brenner and Binford (1988), Deevey et al. (1986), and others. The
importance of assessing land use changes in Florida is further indicated by the high percentage
(57%) of the lakes having evidence of disturbance based on ion chemistry deviations from
expected geochemistry (Pollman and Canfield 1991). Battoe and Lowe (1992) attributed a recent
decline in the pH of Lake Annie in central Florida to acidic deposition. However, preliminary
analyses of aerial photographs show that the watershed has been subjected to numerous land
use changes including construction of extensive ditches that might explain all or part of the
observed changes in acid-base chemistry.

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a.	Monitoring
Historical data on the water chemistry of lakes in the Trail Ridge area of northcentral Florida
have been evaluated by Crisman et al. (1980), Hendry and Brezonik (1984) and Pollman and
Canfield (1991). Analyses by Crisman et al. (1980) and Hendry and Brezonik (1984) were based
on comparison of recent data with data collected by Clark et al. (1964a) and Shannon (1970).
The more recent work by Pollman and Canfield (1991) also included water chemistry data from
ELS-I (Linthurst et al. 1986) and PIRLA-I (Sweets et al. 1990). Of the seven lakes analyzed by
Pollman and Canfield (1991), four showed significant increases in [H+] with time. The other lakes
showed either significant declines (2 lakes) or no trend (1 lake).
The most extensive monitoring data base available was for McCloud Lake, an undeveloped
seepage lake, which had also been sampled from 1980 to 1982 by Baker (1984). The pH of
McCloud Lake decreased about 0.3 pH units from 4.9 in 1968-69 to 4.6 in 1986, but the apparent
trend was driven by Shannon's data collected in 1968. Later surveys (1978-86) suggested short-
term variability, but no consistent trend (Pollman and Canfield 1991).
b.	Paleolimnoloqy
Diatom-inferred pH reconstructions are available for 16 lakes in the Florida Panhandle and
northcentral Florida, nearly all of which lie in upland or ridge regions where soils are deeply
weathered quartz sands and are quite acidic (Carlisle et al. 1978). Paleolimnological
reconstructions of the chemistry of six seepage lakes in Florida were calculated as part of the
PIRLA-I project and reported by Sweets et al. (1990). In addition, 10 Florida seepage lakes were
cored as part of PIRLA-II, and results of these analyses were reported by Sweets (1992).
Paleolimnological study lakes in Florida have been located in the Panhandle, the Trail Ridge Lake
District, and Ocala National Forest, generally in terraces of loose sand (Entisols) that were
deposited on top of the clay confining layer (Hendry and Brezonik 1984). The sands are highly
weathered and soils have very low cation exchange capacity (Carlisle et al. 1981). Many of the

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low-ANC Florida seepage lakes serve as recharge areas for the Floridian aquifer (Baker et al.
1988).
Paleolimnological data collected for Florida lakes in PIRLA-I were not extensive, and
sedimentary evidence for changes in lakewater chemistry in response to atmospheric deposition
was not conclusive. Partial mixing of sediment layers of cores also was common. Of the six
lakes analyzed in PIRLA-I, two (Barco and Suggs) were inferred to have acidified since 1950,
three lakes were inferred to have remained stable or fluctuate with no steady change in pH
(Sweets et al. 1990). The acidification of Lake Barco by 0.3 to 0.8 pH units began about 1950
(Figure IV.C.1). Lake Suggs was inferred to have decreased 0.5 pH units between 1880 and
1920, and a second pH decrease of 0.4 units occurred between 1950 and 1970. The timing of
the onset of inferred acidification after 1950 correlated with increases in S02 emissions and sulfur
deposition, which has been estimated to have increased steadily since about 1945 (Husar et al.
1991). Also, sedimentary accumulation of Pb, Zn, and PAH increased greatly between 1940 and
1950, indicating increased deposition of atmospheric pollutants.
Sweets et al. (1990) provided quantitative estimates of diatom-inferred change in ANC since
pre-1900 for Lakes Barco and Suggs. The diatom-inferred ANC of Lake Barco decreased by
36 fieq L"1 (average of 3 cores) since about 1950, coincident with increases in acidic deposition in
the region. The diatom profile in Lake Suggs suggested two periods of acidification, of
approximately equal magnitude, near the turn of the century and subsequent to 1940. The total
loss of ANC inferred by the diatoms since pre-industrial times at Lake Suggs, was 19 ^eq L"'.
Perhaps half of this change might be attributed to acidic deposition since 1940 (Sweets et al.
1990). If it is assumed that essentially all of the current lakewater concentration of SO.,2' in Lakes
Barco and Suggs is of atmospheric, anthropogenic origin, then approximately 27% of the increase
in S042' in both lakes has caused a stoichiometric decrease in ANC.

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C.1
$
y
j j?
&
-ir ,ir
1980 -
/ yv / -
" ^ / /V / „/ ? *>'
"	.¦&	-»¦' $r ^	s-P	rf> ^
V<*	rf j> J"	v/
¦ "v r A a	F* ^
S^" ^ A^)> 4^4}	^ y
? ^ ^>vv^- ?*\v ^
-I960
-1960
- 1940
-1920 ?
SJ
-1900 3
O
-1880 2
m
i860
CO

1980 -I



«r

1960 -
i


O

1940 -
CJ

1920-
cc
<
C.2
1900 -
CD

1880 -
UJ

I860 -
X
5


N
1980
1960
1900
11920
1900
1880
-i860
C.4
1980 -j
i960 -1
1940
1920
1900
1880
i860
3.5
4.0
I
4.5
—r-
5.0
L ¦ _
L L L
U- i— u
1980
1960
1940
1920
'900
1880
1860
OlATOM INFERRED pH
0 10 0 10 0 10 20 0 5 0 10 0 10 0 10 0 10 OS
PERCENT (--!•»» ttan 1%)
Figure IV.C.1. Diatom-inferred pH and occurrence of major diatom taxa for three cores from Lake
Barco, FL. Abundance-weighted mean pH values for individual taxa are given in
parentheses. Error bars for diatom-inferred pH represent the residual error of the
multiple regression inference equation (Source: Sweets et al. 1990).

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The PIRLA-II project included 10 additional seepage lakes in Florida. Diatom-inferred
changes in pH for the 16 Florida seepage lakes studied to date in both PIRLA projects are
presented in Table IV.C.2. Five lakes were examined in or near the Trail Ridge region, and all
showed some evidence of recent acidification (> 0.2 pH unit decrease; Sweets 1992). With the
exception of Lake Five-O, lakes in the Panhandle region and Ocala National Forest did not show
Table IV.C.2. Summary of diatom inferred pH change from PIRLA cores in north Florida
using maximum likelihood regression (MLR) and the WACALIB
Line and
Birks 1990) inference technique.
Standard errors of prediction:
MLR=0.34
pH units, WACALIB
=0.36 pH units. (Source: Sweets 1993)



MLR
WACALIB



Surface
surface
surface



measured
inferred
inferred
MLR
WACALIB
Lake
PH
PH
pH
change
change:
Trail Ridae





Barco avg.a
4.48
4.17
4.50
-0.69
-0.29
McCloud"
4.65
4.87
4.76
-0.20
-0.13
Magnolia
5.40
5.04
5.41
-0.30
-0.21
Sand Hill (Lowry)
5.32
5.04
5.16
-0.90
-0.93
Suggs6
4.85
5.37
5.51
-0.91
-1.09
Ocala N.F.





Fore"
5.00
5.14
5.11
+0.10
+0;35
Loub
6.33
6.80
7.24
+0.05
+ 019:
Mary5
4.56
4.81
4.67
-0.09
-0:08
Panhandle





Black
5.26
4.63
4.87
+0.05
-0.01
Brock
5.05
5.05
4.85
-0.05
+0.08
Five-Ob
5.04
4.75
4.77
-2.08
-3.00
Lofton
4.69
5.02
4.69
-0.06
-0.01
Major
4.89
5.01
4.80
+0.02
+0.02
Mirrow&
4.87
4.98
4.64
-0.07
-0.01
Moore
4.78
4.96
4.71
-0.01
+0.00
Tommy (Z15)
5.58
5.37
5.05
+0.00
-0.50
a Average of three cores, at least 12 intervals analyzed in each.


6 Cores with at least 12 intervals analyzed. All other lakes had only 4 intervals analyzed.

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evidence of recent acidification. Although Lake Five-0 was inferred to have decreased by 2 pH
units, this acidification cannot be attributed to acidic deposition. The paieolimnological data
suggest that this large decline was associated with a sudden change in lakewater chemistry,
probably caused by a catastrophic disturbance such as sinkhole activity, rather than by
acidification from atmospheric deposition (Pollman and Sweets 1990, Sweets 1992).
c. Process models
Florida is an area with an extensive record of paieolimnological data that can be combined
with information derived from process-based models. A generally reasonable agreement between
the inferred and modelled hindcasts is illustrated for Barco Lake (Figure IV.C.1), located in
northcentral Florida, although the model estimated somewhat higher pre-industrial pH, and
therefore greater acidification than did the diatom approach.
Pollman and Sweets (1990) conducted hindcast simulations for 15 lakes in the Florida
Panhandle and northcentral Florida, using the IAG model. Study lakes were those included within
the PIRLA-I and PIRLA-II paieolimnological investigations. Model hindcast projections were
constructed assuming a range of base cation neutralization of S042' in groundwater from 0%
(F=0) to 100% (F=1.0). IAG model results were compared with diatom inferences of historical
change in pH. Although Pollman and Sweets (1990) concluded generally good agreement
between the two approaches, this agreement was primarily confined to those lakes inferred by
both techniques to have not experienced acidification. Those lakes inferred by either technique to
have experienced acidification of more than the standard error of the diatom inference equation
(0.31 pH units) showed generally poor agreement between the IAG model and the diatom
hindcast approaches (Figure IV.C.2).
As pointed out by Pollman and Sweets (1990), there are a number of assumptions required
for conducting IAG hindcast simulations, and these assumptions can have very large influence on
model output. They include:

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-2.5 -2 -1.5 -1 -0.5 0 0.5 1
Diatom-inferred change in pH since pre-industrialization
Figure IV.C.2. Comparison of diatom-inferred and IAG model hindcasts of pH change from pre-
industrial times to the present for fourteen lakes in northern Florida. IAG model
scenarios were conducted assuming groundwater concentration factors of 1.25
and base cation neutralization of sulfate in groundwater ranging from 0% (F=0) to
100% (F=1.0). Data from Lake Five-O, which appears to have experienced major
hydrologic changes, was deleted from the analysis. (Data from Pollman and
Sweets 1990.)

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•	assumed input trajectories for sulfur and nitrogen,
•	assumed input trajectories for base cations, and
•	neutralization effects of DOC
Assumptions regarding base cation deposition are more important for IAG modeling of seepage
lake chemistry than for modeling drainage systems, using a model such as MAGIC (e.g., Sullivan
et al. 1990b). Pollman et al. (1990) found IAG hindcast simulations to be extremely sensitive to
assumed pre-industrial base cation deposition. For example, changing this estimate by 10%
resulted in factor-of-two differences in estimates of historical acidification for Lake Barco. Thus,
although modeling studies may aid in improving our understanding of the response of seepage
lakes to acidic deposition, basic geochemical and hydrological data, with which to parameterize
the models, are generally lacking.
D. EASTERN CANADA
1. Characteristics and Sensitivity to Acid Deposition
The sensitivity of aquatic resources in eastern Canada to acidic deposition has been inferred
from a number of independent regional surveys which were reviewed and synthesized by Cook et
al. (1988) and Jeffries (1991). Both of these reports relied heavily on the lake database compiled
and described by Jeffries (1986). As has been the case in much of the United States, most of the
Canadian surveys have focussed on lakes. We are unaware of data suitable for comparing the
sensitivity of streams versus lakes in eastern Canada. The area of interest for defining sensitive
resources in eastern Canada used by Cook et al. (1988) included the provinces east of Ontario
and south of latitude 52° N, although data used in both Cook et al. (1988) and Jeffries (1991) also
included lakes in Labrador extending north of 52° N. The areas considered sensitive to acidic
deposition were delineated by Shilts (1981). Precambrian bedrock and glacial till with poorly
weatherable minerals extends across much of southeastern Canada, encompassing over 40% of
the land area (Jeffries 1991). Soils in the region are generally coarse textured and low in

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exchangeable base cations. Most of the lakes are drainage lakes, due to the presence of thin
glacial deposits over silicate bedrock and annual precipitation generally > 100 cm/yr (Jeffries
1991).
2. Current and Projected Future Deposition
Emissions of S02 and NOx for Canada were estimated at 3.7 Tg/yr and 1.9 Tg/yr, respectively
in 1985, which represented 17.5% and 10.1% of S02 and NO, emissions in the United States.
The S02 emissions in Canada during 1985 had declined 45% from emissions during 1970. Nearly
all of the decrease in S02 emissions was associated with declines from industrial sources (Placet
et al. 1990). Husar et al. (1991) estimated historical S02 emissions trajectories for southeastern
Canada and observed that emissions increased dramatically in the 1930's and agajn in the
1960's. Major decreases in S02 emissions that occurred in the U.S. during the 1930's from the
economic depression and during the 1945-1950 period following WW-II were not evident in the
data on Canadian S02 emissions. NOx emissions in Canada have increased about 10% since
1970, with most of the increase attributable to external fuel combustion and industrial processes.
Ammonia emissions in 1985 were estimated at 0.2 Tg/yr, of which three-fourths were generated
by animal waste and fertilizer applications (Placet et al. 1990).
Excluding areas adjacent to local sources, wet sulfur deposition is greatest in southeastern
Canada, reaching 12 kg S ha1 yr1 and decreasing to background levels (~ 2 kg S ha'1 yr'1) in
Newfoundland and extending westward to western Ontario (Jeffries 1991). Wet nitrate plus
ammonium deposition ranged from about 7 to 2 kg N ha1 yr'1 (Sisterson et al. 1990). Annual
composite precipitation pH for 1980-1984 in southeastern Canada ranged from 4.2 to 4.6
(adopted from Barchet 1987).

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3. Current Surface Water Chemistry
The chemistry of lakes in southeastern Canada has been surmised on the basis of numerous
independent surveys of over 7400 lakes (Jeffries 1991). Consequently, the various investigators
did not determine a common list of analytes or employ common methods. The most commonly
measured parameters in these surveys were pH, ANC, Ca2+, Mg2+, and S042', which are
summarized for eight subregions in eastern Canada (Table IV.D.1). Most of the Ontario lakes
appear to be well buffered as evidenced by high concentrations of base cations in all three
subregions of Ontario. Jeffries (1991) attributed this to the irregular presence of carbonate in the
glacial overburden. Despite these high base cation concentrations, the small percentage of the
study lakes that were acidic likely represents a large number of acidic lakes in the population.
For example, Kelso et al. (1986) estimated that 14,000 lakes were acidic (pH < 4.7) in
southeastern Canada, whereas Jones et al. (1984) estimated 12,500 lakes had pH < 5. Jeffries
(1991) believed that these estimates were conservative because many of the smaller, and
presumably more acidic lakes, were not well represented in the surveys and because most of the
data were collected in summer and fall when effects of episodes would not be present.
Furthermore, Cook et al. (1988) noted that biological effects of acidification would be expressed
well before the lakes had decreased to pH 5.0 (ANC ~ -10 fieq L'1).
Region-wide data are not available regarding concentrations of organic anions in
southeastern Canada. In the Atlantic provinces of Nova Scotia, Newfoundland, and Labrador,
extensive peat deposits generate considerable organic anions (A ), but in Labrador where many
lakes have high A', most lakes have pH > 6.0 (Jeffries 1991). Some of the highest concentrations
of A" among Canadian lakes are located in northwestern Ontario where sulfate deposition is low.
Both Jeffries (1986) and Neary and Dillon (1988) concluded that there was little relationship
between the distribution of A' and low-ANC lakes in Canada.
Sulfate contributions from watershed 'sources are important in some areas of the United
States (e.g., Minnesota, West), but were considered unimportant in most areas of southeast
March, 1994
Page 88

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Table IV.D.1. Quarlile pH and Ca2+ + Mg2+, ANC, and S042 concentrations (fieq L"1) for (he eight subregions of eastern Canada.
(Source: Jeffries 1991)
Ion
Percentile
NF*b
NSab
NBbc
LAB"b
QUE'
NW ONT'
NE ONT*
SC ONT*
PH
25th
5.0
5.0
5.8
6.1
5.8
6.7
6.3
5.9
50th
6.2
5.4
6.3
6.4
6.2
7.1
6.9
6.3

75th
6.6
6.0
6.9
6.6
6.6
7.5
7.4
6.7
Ca2* + Mg24
25th
62
61
70
54
118
173
224
164

25lhc
43
31
55
48





50 th
98
65
115
73
163
235
383
194

50thc
79
53
90
69





75th
141
135
164
100
234
477
1008
236

75thc
118
84
156
100




ANC
25 th
12
-2
6
28
26
129
53
30

50 th
28
14
38
46
52
261
166
61

75th
50
38
83
82
112
622
615
121
(Nl
o
CO
25th
29
50
65
19
62
62
94
141

25thc
21
36
60
16





50th
39
71
83
26
94
75
142
156

50thc
31
50
77
24





75th
50
85
94
31
135
94
185
172

75thc
40
65
89
30




* Subregion abbreviations: NF = insular Newfoundland, NS = Nova Scotia, NB = New Brunswick, LAB = Labrador, QUE = Quebec.
NW ONT - northwestern Ontario, NE ONT = northeastern Ontario, and SC ONT = southcentral Ontario
b Sea salt corrected data are presented for only NF, NS, NB, and LAB; input of marine salts to the remaining subregions is
inconsequential
c Sea salt corrected Ca2+ + Mg2* and S042 are given for those subregions experiencing significant deposition of marine salts
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March, 1994
Page 90
Canada (Jeffries 1991). Cook et al. (1988) concluded that most of the S042" concentrations in
Canadian lakes could be attributed to atmospheric deposition.
4. Quantitative Assessment of Acidification
a. Monitoring
Comparisons of chemical data 1rom surface waters sampled both in the past and more
recently in Canada were summarized by MOI (1983) and Cook et al. (1988). Comparisons for
Nova Scotia rivers sampled in 1954-55 and 1980-81 (Watt et al. 1983) and Newfoundland rivers
sampled over the period 1960 to 1980 (Thompson et al. 1980) suggested recent acidification
and/or recovery consistent with acidic deposition effects. Wiltshire and Machell (1981) reported
pH and ANC declines for 10 remote lakes in Nova Scotia and New Brunswick sampled in 1950
and 1979. The conclusions of these studies are, however, based on historic measurements
having a large uncertainty that may obscure the magnitude of the pH trends (Cook et al. 1988).
Accurate and reliable measurements of lakewater chemistry have been made for lakes in the
Sudbury area of Ontario since about the mid 1970s. Nickel and copper containing sulfide ores
• have been mined and smelted near Sudbury throughout this century. These activities have
resulted in substantial acidification and trace metal contamination of nearby lakes and elimination
of adjacent forest vegetation. The construction of a 381-m stack in 1972 caused greater
dispersion of S02 emissions, and concurrently total emissions were also reduced. S02 emissions
at Sudbury decreased substantially after 1970 (Dillon et al. 1986), and lakewater chemistry
response has been well documented (Dillon et al. 1986; Hutchinson and Havas 1986; Keller and
Pitblado 1986; Keller et al. 1986). Sulfate concentration in Clearwater Lake decreased from an
average of 545 peq L"' during the period 1973-77 to 370 ,ueq L'1 in 1984. The decrease of 175
,ueq L' of S042" during one decade is considerably greater than'the S042' enrichment that has
occurred in Adirondack lakes from pre-industrial time to the present. Lakewater pH increased
from 4.2 to 4.6 and total Al decreased from 430 to 190 ng/L in Clearwater Lake during the study

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period (Dillon et al. 1986). In Swan Lake, the decrease in [S042] was even more pronounced,
from 580 to 220 ^eq L'1 between 1977 and 1982, and pH increased from 4.0 to 4.8. After
correcting for base cations that were charge balanced by CI', CB decreased from 441 ^eq L"' to
212 fieq L'\ and the calculated F-factor was 0.64 (Dillon et al. 1986). Cook et al. (1988)
suggested that the calculated F-factor for Swan Lake was unusually high because this lake has
very high CB for an acidic lake. It is more likely, however, that it is precisely because the F-factor
is so high that the base cation sum is also high. In Baby Lake, near Sudbury, the decrease in
[S042] has been extremely large (-750 fieq L1) and pH has increased from 4.0 in 1972 to 5.8 in
1984 (Hutchinson and Havas 1986). Sulfate concentrations in Laundrie Lake and Florence Lake
(200-300 jueq L"') were closer to the concentrations observed in areas of the United States that
receive acidic deposition. S042" concentrations decreased by about 40 to 60 fieq L"1 in these
lakes during the period 1974-1976 to 1979-1983. The average changes in [H+] relative to the
change in [S042] for these two lakes during mis period were 0.24 and 0.22, respectively (Keller et
al. 1986).
During 1981-1983, 209 lakes in the vicinity of Sudbury that had been sampled in 1974-76
were resampled (Keller and Pitblado 1986; Keller et al. 1986). Significant decreases in lakewater
S042' concentration and increases in pH were observed. pH recovery was generally limited,
however, to lakes with pH < 6.0 (Figure IV.D.1). The magnitude of pH recovery was
associated with distance from Sudbury; lakes in close proximity to the smelters exhibited the
largest response. It was not clear, however, whether the pattern depicted in Figure IV.D.1 was
due to the magnitude of [S042] change, responsiveness of the systems, or both. It is significant,
however, that paleolimnological research in most regions generally indicates that lakes currently
having pH greater than about 6.0 have not experienced historical decreases in pH or ANC in
response to acidic deposition, and similarly, only the Sudbury lakes with pH < 6.0 showed
evidence of recovery (Sullivan 1990).

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i
«¦» 7
• <

O. •
z
<
UJ
a
s
4
A
s
MEAN PM 1974-7 9
7
Figure IV.D.1. Average pH in 1974-76 versus average pH in 1981-83 for lakes in the vicinity of
Sudbury, Ontario. The line represents a 1:1 relationship. Filled circles indicate
coincident points. (Source: Cook et al. 1988, redrawn from Keller and Pitblado
1986).
Kelso and Jeffries (1988) reported changes in lakewater chemistry between 1979 and 1985
for 54 lakes in the Algoma region of Ontario, located approximately 200 km from Sudbury. The
annual precipitation-weighted SO.,2" concentration of wetfall exhibited a small downward trend
during the period 1976 to 1985. Sulfate deposition was lower after 1980 in response to both
lower precipitation quantities and lower average [SO/ ] in wetfall. Lakewater chemistry data were
reported for three sampling periods, 1979, 1982, and 1985. Median lakewater [SO„2'] decreased
during the period from 133 to 106 ueq L'1. pH increased in nearly all lakes between 1979 and
1982, but showed little change between 1982 and 1985. Median pH for the three sampling
periods were 6.35. 6.64, and 6.65, respectively. These overall increases in pH did not agree with

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median ANC values, however, which were 79, 86, and 76 fieq L'1. The reported increases in pH
during the study period for lakes with pH > 6.0 should be interpreted with caution, since they do
not agree with reported ANC changes (which should be more responsive than pH to acidification
or recovery in this range). Only in 1979 were any lakes sampled with negative ANC or pH < 5.0.
The minimum lakewater pH increased during the study period from 4.60 to 5.44.
Kelso and Jeffries (1988) calculated the change in [Ca2* + Mg2*] relative to the change in
[S042] for the lakes in their data set and obtained an estimated median F-factor of 0.79. Many of
the lakes that were lowest in [Ca2+] (minimum = 71 «eq L"') were calculated to have F-factors in
the 0.1 to 0.4 range, however.
Gunn and Keller (1990) sampled 67% of all known lakes with lake trout populations (ranging
from extinct to healthy populations) within a 1.4 x 103 km2 area around Sudbury. Water samples
were collected in January 1980 and January 1987. The average increase in pH between sample
occasions was 0.37 pH units, and the average increase in ANC was 23 ^eq L'1. Sulfate
concentrations decreased, on average, 45 ^eq L1 during this period (J. Gunn, personal
communication). Thus, assuming no large changes in N03' or Al concentrations, the average F-
factor for the recovery of these lakes was about 0.49. Gunn and Keller (1990) also reported
annual means, based on monthly samples, of selected water quality parameters for Whitepine
Lake, a headwater lake in a forested catchment 90 km north of Sudbury. Lakewater S042"
concentration declined from 237 ^eq L' in 1980 to 195 /xeq L'1 in 1988, an 18% decrease. Sulfur
emissions at Sudbury declined during the same period by 23%, from 935 to 718 thousand tonnes
per year. Approximately 24% of the decrease in SO„2" concentration in lakewater was balanced
by an increase in ANC, and the lakewater pH increased from 5.4 to 5.9.
In central Ontario, atmospheric sulfur deposition declined about 35 to 40% since the mid
1970's: the annual average concentration of S042 in precipitation declined about 4 /ieq L"' yr1
over the period 1976 to 1985 (Dillon et al. 1988). Despite the decline in sulfur deposition,
however, lakewater chemistry monitoring at Plastic Lake, a small dilute lake on the Precambrian

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Shield, showed acidification between 1979 and 1985 (Dillon et al. 1987). Alkalinity declined an
average of 2 fieq L' yr'\ largely in response to changes in base cation concentrations.
Deposition of sulfur remained relatively constant since 1986, and concurrent lakewater chemistry
was reported by Dillon and LaZerte (1992). Water quality of runoff draining a small upland site
improved from 1983 to 1987; the concentrations of S042' and Al decreased, whereas the pH and
alkalinity increased. Sulfate concentration at the upland site increased abruptly in 1988, as did
the base cation concentrations, and alkalinity decreased (Dillon and LaZerte 1992). Thus, the
acid-base chemistry of the upland site was not closely coupled with changes in deposition during
the study period. Changes in water chemistry at the base of the catchment in the Plastic Lake
inflow, below a wetland area, were very different from changes in chemistry at the upland site.
Sulfur dynamics at the inflow were strongly influenced by climate. Subsequent to dry periods in
1983, 1987, 1988, and 1989, S042' concentrations increased to very high levels due to re-
oxidation of stored sulfur in the wetland. These periodic excursions of high SO/' concentration
were accompanied by increases in base cation concentrations and decreases in alkalinity.
Longer trends in acid-base chemistry that were observed at the upland site were not apparent
below the wetland (Dillon and LaZerte 1992).
Only a limited number of studies have been conducted of episodic acidification in eastern
Canada. Wigington et al. (1990) summarized the results of these studies (n=10), which have
been conducted primarily in Ontario and Quebec. Most of the research efforts have focussed on
snowmelt events. Three lakes (Dickie, Harp, Plastic) in the Muskoka/Haliburton region have been
intensively studied by Dillon, Jeffries, and co-workers (e.g., Jeffries et al. 1979, Dillon and Molot
1989, Molot et al. 1989). Snowmelt events cause episodic pH depressions in Dickie Lake and
Harp Lake of about 0.8 to 0.9 pH units for periods up to 15 to 30 days (Jeffries et al. 1979).
Dilution of base cations is the principal process responsible for ANC depression, although
increases in N03" and SO„2' concentration can augment the dilution effect (Dillon and Molot 1989,
Molot et al. 1989). Snowmelt events in the early part of the melting season are caused mainly by.

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peaks in N03" concentration at Plastic Lake. pH can be depressed by 0.7 pH units (to pH 5.0) in
the littoral zone, whereas in deeper water, pH declines in a more gradual fashion to about 5.4
(France and LaZerte 1987).
At the Turkey Lakes watershed in Ontario (close to the border with Upper Peninsula of
Michigan), dilution of base cations appears to be the principal cause of episodic ANC
depressions (Jeffries and Semkin 1983). Lakewater pH declines by about 0.5 to 1.0 pH unit
during episodes, although bottom water shows smaller variation (Jeffries et al. 1988). The most
extensive studies of episodic acidification in Quebec have been conducted at Lac Laflamme in the
Laurentian Mountains. Snowmelt results in decreases in ANC of about 200 fieq L"\ and pH
declines by 1.0 to 1.5 pH units to minimum values of 5.0 to 5.4. These episodic changes are due
primarily to base cation dilution and increase N03 concentration (Wigington et al. 1990).
b. Paleolimnoloav
A large percentage of Canadian paleolimnoiogical studies have focused on lakes near large
point sources (e.g., Sudbury), near mining effluents, or on lakes that have been manipulated
(limed or acidified). In addition, several studies included lakes strongly influenced by organic rich
soils and bogs (Charles et al. 1989). In these organically-dominated systems, pH changes in
response to acidic deposition would not be expected to be large because of organic buffering.
Thus, a regional paleolimnoiogical data base with which to evaluate acidification associated with
long-range transported pollutants is not available in Canada. In regions where acidic precipitation
occurs, however, recent pH declines have been inferred for at least some lakes.
Paleolimnoiogical data have been published for several lakes in eastern Canada (e.g.
Delorme et al. 1984; Dickman and Fortescue 1984; Dixit 1986; Dixit et al. 1987, 1988, 1989a,
1989b). Kejimkujik Lake in central Nova Scotia is currently acidic (pH 4.8) with high DOC
concentrations. Paleolimnoiogical data suggest that this lake has been acidic for the past 1000

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years, presumably due to organic acidity, and that the lake has not undergone recent acidification
(Delorme et al. 1984; Charles et al. 1989).
Acidification has been inferred for many lakes located near smelters in Wawa and Sudbury,
Ontario (Dickman and Fortescue 1984, Dixit 1986, Dixit et al. 1992d). Most lakes in areas distant
from local point sources (Ottawa Valley, southwest Quebec, Laurentide, northern Quebec) were
not inferred to have acidified (Cook et al. 1988, 1989). Most of these lakes have pH values above
6.0 and are probably insensitive to acidic deposition effects.
Cook et al. (1989) compiled paleolimnological data on Canadian lakes from the scientific
literature (Charles et al. 1989) and unpublished data from J.P. Smol, Queen's University. The
Canadian data set used in the Charles et al. (1989) review included data for 81 lakes, and Smol
provided data for 5 additional lakes. Of the lakes examined by Cook et al. (1989),
paleolimnological data for 48 were in the peer-reviewed literature, including diatom-inferred past .
and present pH values, thus allowing estimation of pH change (Table IV.D.2). The study lakes
are not, however, representative of lake populations in eastern Canada. The greatest changes in
diatom-inferred pH occurred in lakes proximate to the Sudbury (median change, -1.3 pH units)
and Wawa (median change, -0.4 pH units) smelting and scintering plants, and in Lake 223, which
was experimentally acidified (Table IV.D.2). .
The North and South Atlantic Regions exhibited median changes in inferred pH for the study
lakes of -0.1 and -0.2 pH units. The year of first change was 1920 to 1950 for the lakes exhibiting
evidence of acidification (Table IV.D.2). These regions receive high rates of acidic deposition, and
of the lakes sampled for lakewater chemistry, 12% and 3%, respectively, have ANC < 0 fieq L'1
(Baker et al. 1990). The acidification of these lakes is likely attributable to atmospheric deposition
as well as possible watershed disturbances (Table IV.D.2).
Southeast Ontario (Region 3), the region receiving the highest rates of atmospheric
deposition and having a large number of sampled lakes with ANC < 0 ueq L"' (86 out of 2080
lakes), had only two lakes for which diatom-inferred pH changes could be estimated. One of the

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Table IV.D.2. pH of lakes in eastern Canada inferred from paleolimnological data (Source: Cook et al.
, 1989)








Current

Year of





Diatom Inferred pH
Measured
Cause of
First

Region'
Province
Lake Name
Method2
Past
Current
Delta
PH
Change3
Change4
Reference5
1
Ontario
Lake 223
1
6.7
5.4
-1.3
5.1
5
1976
1
3
Ontario
Batchawana L.
2
6.0
6.2
0.0
6.1
4
NA
2

Ontario
Quirke L.
2
6.2
5 5
-0.7
6.0
2
1957
3
4
Quebec
David (P.Q.)
3
7.0
7.1
0.0
7.3
NA
NA
4
5
Quebec
Blais
3
7.3
7.2
0.0
7.1
NA
NA
4

Quebec
Chevreuil
3
7.0
7 1
0.0
7.0
NA
NA
4

Quebec
Kidney
3
6.8
6.8
0.0
6.8
NA
NA
4

Quebec
Truite Rouge
3
5.8
5.7
0.0
6.1
NA
NA
4
6
Quebec
Bonneville
3
5.6
5.4
-0.2
5.4
NA
1940
4.5

Quebec
Chomeur
3
7.0
6.9
0.0
6.7
NA
NA
4

Quebec
Lagou
3
5.7
5.6
0.0
5.8
NA
NA
4

Quebec
Nolette
3
6.5
6.4
0.0
6.5
NA
NA
4

Quebec
Thomas
3
6.5
6.4
0.0
6.8
NA
NA
4
7
Quebec
Lemaine
3
6.1
6.2
0.0
6.0
NA
NA
4
8
N. B.
Emigrant L.
1
6.5
6.1
-0.4
6.1
1
1935
6 ,

N. B.
Lilly L.
1
6.2
5.7
-0.5
5.6
1
DU
6

N. B.
St. Patrick's L.
1
6.8
6.8
0.0
6.7
NA
NA
6

N. B.
Tamoowa L.
1
6.4
6.4
0.0
5.3-
NA
NA
6

N. S.
Big Indian L.
1
6.1
5.3
-0.8

1
1940
6

N. S.
Kejimkujik L.

4.7
4.7
0.0
4.8
NA
NA
7.8

N. S.
Kinsac L.
1
6.3
6.1
-0.2
5.5
1
1920
6

N. S.
Round L.
1
6.2
6.2
0.0
4.6
NA
NA
6










Continued
ft)
3
a
C?
O
iO
ffl
3
33
©
XX
c
&
s
5 2

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Table IV.D.2. Continued







Current

Year of





Diatom Inferred dH
Measured
Cause of
First

Region1
Province
Lake Name
Method2
Past
Current
Delta
PH
Change3
Change4
Reference5
9
Newf.
Aides P.
3
6.0
6.4
0.4 *
6.3
NA
1930
9

Newt.
Stephensons
2
6.0
5.9
0.0
5.6
NA
NA
9

Newt
12
3
5.4
5.1
-0.3
5.1
1?
DU
9,10

Newf.
204
3
5.5
5.2
-0.3
5.1
1?
DU
9,10

Newf.
219
3
5.5
5.5
0.0
5.6
NA
NA
9,10

Newf.
660
3
5.7
5.5
-0.2
5.5
1?
1953
9,10

Newf
8
3
5.5
5.3
-0.2
5.3
NA
DU
9

Newf.
817
3
5.7
5.3
-0.4
5.2
1?
1930
10,11

Quebec
C-22
1
5.5
5.4
0.0
5.8
NA
NA
12

Quebec
Key
1
5.3
5.2
0.0
5.5
NA
NA
12
10
Ontario
Acid
4,6
6.0
4.4
-1.6

2
1920
13

Ontario
Baby
4,6
6.5
4.4
-2.1

2
1920
13

Ontario
Clearwater
3,5
6.0
4.2
-1.8
4.5
2
1930
14

Ontario
Hannah
3.5
6.0
4.6
-1.4
4.3
2
1880
14

Ontario
Labelle
4,6
6.2
6.2
0.0
6.2
NA
NA
13

Ontario
Lake 29 B
2
5.3
4.0
-1.3

2
1970
17

Ontario
Lohi
6
6.1
4.8
-1.3
4.7
2
1940
14,16

Ontario
Swan
7
5.8
.3.8
-1.9

2
1940
16

Ontario
White Pine
2
5.9
5.8
0.0

NA
NA
17
11
Ontario
B
1
6.4
4.7
-1.7
5.2
2
1955
18

Ontario
Beaver L.
1
5.7

-0.5
5.2
2
1954
19

Ontario
CS
1
7.2
6.4
-1.0
5.2
2
1955
19

Ontario
Fenton L.
1
7.2
7.1
0.0

2
NA
19

Ontario
U1
1
6.3
6.3
0.0
6.4
NA
NA
20

Ontario
U4
1
6.6
7.1
0.5

NA
NA
20

Ontario
W1
1
7.0
7.2
0.0
7.1
NA
NA
20
m >
w o g
III
-i 2T O
§ 3 3.
3 > 2.
a ¦§ o
sr & ®
_ eg
n"
5
i. 8 §
6	c »-
•3 8 5
- 2 ®
3 (A
W
C
fil
3
a
o
to

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Table IV.D.2. Continued
1	Regions: 1-West Ontario, 3-Southeast Ontario, 4-Ottawa Valley, 5-Southwest Quebec, 6-Laurentide, 7-North Quebec, 8-South
Atlantic, 9-North Atlantic, 10-Sudbury, 11-Wawa.
2	Methods: 1-diatoms, index a\ 2-diatoms, index B; 3-diatoms, multiple linear regression; 4-diatoms, canonical analysis; 5-
chrysophytes, qualitative; 6-chrysophytes, canonical analysis; 7-chrysophytes, multiple linear regressions.
3	Cause of change: 1 -acid deposition; 2-local acid deposition; 3-watershed disturbance; 4-climate; 5-experimental acidification; NA-
not applicable.
4	Date of first change: DU-core not dated; NA-not applicable.
5	1 -Dickman et al. 1988; 2-Delorme et al. 1986; 3-McKee et al. 1987; 4-Dixit et al. 1987; 5-Dixit and Dixit 1989; 6-Elner and Ray 1987;
7-Delorme et al. 1984; 8-Duthie 1989; 9-Scruton et al. 1987a; 10-Scruton et al. 1987b; 11-Rybah et al. 1989; 12-Hudon et al. 1986;
13-Dixit and Dixit, pers. comm.; 14-Dixit and Evans 1986; 15-Dickman and Rao 1988; 16-Dixit et al. 1989; 17-Dickman and Rao
1989; 18-Dickman et al. 1985; 19 Dickman and Thode 1985; 20-Dickman and Fortescue 1984.

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two lakes exhibited no change in diatom-inferred pH. This lake is located in an area (north of
northern Michigan) that receives low levels of atmospheric deposition compared to the southeast
Ontario region as a whole. The acidification of the second lake was attributed to local mining
andmilling of uranium (McKee et al. 1987). Following cessation of mining and milling in the
1960s, the lake recovered to pH 6.4.
Lakes in Canada that have experienced acidification, based on paleolimnology, generally
have current pH < 6.0. These lakes have insufficient buffer capacity in their watersheds to fully
neutralize acidic atmospheric inputs. Lakes with pH > 6.0 (except some near Sudbury) appear
not to have experienced chronic acidification, likely because the watershed-lake systems are
sufficiently buffered. These results are in agreement with a statistically-based paleolimnological
assessment of lakewater acidification in the Adirondacks, in which diatom-inferred acidification
was generally restricted to lakes having pH < 6.0 (Sullivan et al. 1990a).
Dixit and co-workers have reconstructed the pH histories of over 70 lakes in the general
vicinity of Sudbury, documenting both the acidification and recovery responses of nearby lakes to
smelter operations (e.g., Dixit et al. 1987; 1989a,b; 1991; 1992a,b,d). Diatom and chrysophyte
inferences clearly reflect the known acidification chronologies for lakes in this region, many of
which experienced very large pH changes in response to changing local sulfur emissions. An
inferred pH decline of 2 units was reported for Chiniguchi Lake, located 56 km NE of Sudbury, by
Dixit et al. (1991). Hannah and Baby lakes acidified by about 1.5 to 2 pH units, with minimum pH
values in the 1970s. Closure of the nearby Conniston smelter in 1972 was reflected in a pH
recovery to pH 6.2 in 1987 in Baby Lake (Dixit et al. 1992b). Hannah Lake was limed in 1975.
The pH rise in both lakes was reflected in their diatom inferences (Dixit et al. 1992b).
Dixit et al. (1992d) analyzed surface (recent) and bottom (pre-1880) sediment samples of 72
lakes within 100 km of Sudbury for diatom valves and chrysophyte scales. The extent of
acidification (A pH) inferred from the diatoms varied according to proximity of the lakes to local
smelters, orientation of prevailing wind patterns, and differences in watershed geology.- The most

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extensive acidification was reported for lakes having current pH < 6.0. These were inferred to
have acidified by 0.01 to 1.79 pH units. Some lakes having current pH 6.0 to 7.0 and all lakes
having current pH > 7.0 were inferred to have become more alkaline since pre-industrial times
(Dixit et al. 1992d). This pattern was in general agreement with results of a comparable study in
the Adirondacks (Sullivan et al. 1990a, Cumming et al. 1992).
Results of paleolimnological reconstructions obtained by Dixit et al. (1992d) suggested that
the percentage of Sudbury area lakes having pH < 5.5 has increased from 8 to 29%, and 27% of
the lakes were inferred to have had pH < 6.0 during pre-industrial times. This suggests
that current Ontario Ministry of the Environment guidelines, which characterize lakes as
"damaged" if their pH is below 6.0 (Neary et al. 1990), are not realistic (Dixit et al. 1992d).
Dixit et al. (1992a) reconstructed the pH responses of three lakes located in the La Cloche
Mountains, 60 km SW of Sudbury. Background pH values for the three lakes were in the range
5.4 to 5.8. Diatom-inferred pH declines for the lakes ranged from about 0.4 to 0.7 pH units. All
three lakes showed some increase in pH by the late 1980s, the most pronounced of which was
for Lake Lumsden which recovered from pH 4.7 to 5.3 (close to background, pre-acidification
values) (Dixit et al. 1992a).
c. Experimental Manipulation
Lake 223 in the Experimental Lakes Area (ELA) in northwestern Ontario has been the subject
of an ecosystem-scale acidification experiment. Over a period of 8 years, the pH of the lake was
gradually lowered from 6.8 to 5.0 by the addition of sulfuric acid. Changes in physical, chemical,
and biological factors were reported by Schindler and Turner (1982), Schindler et al. (1985), and
Cook et al. (1986). A whole-lake alkalinity and ion budget for the lake showed that 66% to 81% of
the added sulfuric acid was neutralized by alkalinity production within the lake (Cook et al. 1986).
Based on calculations of changes in ion storage within the lake, Cook et al. (1986) estimated that
the base cation change relative to the [S042 ] change during the first 8 years of manipulation was

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0.4. This value probably underestimates base cation neutralization because it is based upon
acidification of the lake only. Potential increase in base cation mobilization from the catchment is
not included in calculations from a lake acidification experiment, such as at Lake 223.
E. COMPARISONS WITH DATA FROM THE ADIRONDACK MOUNTAINS
As described in the previous sections, quantification of sulfur and nitrogen dose-response
relationships is difficult for the regions considered in this report (Western and Upper Midwestern
United States, Florida, and Eastern Canada). Limited data availability precludes rigorous
quantitative assessment in most cases. In particular, data are scarce in the following categories:
•	episodic acidification, especially in the West
•	groundwater inflow (and associated neutralization) to seepage lakes
•	seasonal surface water chemistry data, particularly for nitrogen and aluminum
. • model input parameters (especially soils characteristics) for drainage systems
•	deposition (wet and dry) data at high elevation sites
•	regional paleolimnologicai data, especially in upper Michigan.and portions of eastern
Canada
¦ The data that are available for the regions under investigation for this report can, however, also be
compared and contrasted with more intensive data generated in other regions that have been
impacted by acidic deposition and studied in greater detail. Such comparisons.can be useful to
place bounds on the magnitude of the acidification response. Quantitative dose-response
relationships for sulfur have been determined, using a variety of approaches, in a number of
regions in North America and Europe.
The Adirondack region of New York is the most intensively studied area in the United States
where multiple independent sources of data are available with which to quantify past and future
regional changes in surface water chemistry in response to acidic deposition. Available data
include MAGIC model reconstructions of past (Wright et al. 1986, Sullivan et al. 1992) and future
March, 1994
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(Church et al. 1989, Thornton et al. 1990, NAPAP 1990) acid-base chemistry; paleolimnological
reconstructions of pre-industrial chemistry (Sullivan et al. 1990a, Charles and Smol 1990); recent
survey and monitoring data (Linthurst et al. 1986, Kretser et al. 1989, Driscoll et ai. 1991, Driscoll
and van Dreason 1993); and historical survey data (Schofieid 1976, Kramer et al. 1986, Asbury et
al. 1989). MAGIC simulations of future changes in chemistry (NAPAP 1991) were based on
projections of constant, increased (+20%, + 30%), and decreased (-20%, -30%, -50%) deposition
over the next 50 years.
Although substantial variability was found in projected future change in ANC among the
modeled Adirondack watersheds, Sullivan et al. (1992) found a highly consistent relationship
40
30
20
10
0
10
•20
•30
•40
¦3
¦6
•4
¦2
0
2
4
6
Meoian Charge in Sulfur Deposition (kg/hatyr)
Figure IV.E.1. Median and range of projected change in ANC (^eq L"') of Adirondack lakes for
50-year MAGIC simulations versus median change in sulfur deposition (kg ha'1 yr'1)
for each deposition scenario (points on each line correspond to -50%, -30%, -20%,
0%. +20%, +30% change from current deposition). (Source: Sullivan et al. 1992)

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between median change in acidic deposition and projected median change in ANC over 50 years.
Each 1 kg ha'1 yr'1 change in sulfur deposition caused approximately a 3.5 ^eq L"' change in
simulated lakewater ANC (Figure IV.E.1). The median (and range) of estimated change in ANC
was +13 fieq L"1 (6 to 34) for the -50% deposition scenario and -15 [teq L"1 (-26 to -8) for the
+30% deposition scenario (Sullivan et al. 1992).
Results of MAGIC model hindcast simulations suggested that all of the modeled Adirondack
lakes had acidified (decreased in pH or ANC) since pre-industrial times. The median and range
of estimated changes in ANC were -46 /zeq L1 and -31 to -84 /zeq L'\ respectively. None of the
lakes were inferred to have been acidic in pre-industrial times. The minimum simulated pre-
industrial values were pH 6.1 and ANC = 30 /zeq L'1 (Sullivan et al. 1992). Upon inclusion of an
organic acid representation in the MAGIC hindcast simulations, estimates of pre-industrial pH
decreased substantially, and the minimum estimated pre-industrial pH was 5.4.
Diatom inferences of historical changes in Adirondack lakewater chemistry suggested the
following:
•	the "median" Adirondack lake has not acidified
•	acidification was generally limited to lakes that have current Gran ANC less than about
50 ^eq L'1 (or pH about 6.0)
•	approximately 15% of the Adirondack lakes were inferred to have acidified by more than
0.28 pH units (apparent RMSE of diatom inference equation)
•	the median historical acidification (expressed as A ANC - A AIJ of currently acidic lakes
was -37 fieq L'1
•	approximately 3% of the Adirondack lakes were acidic in pre-industrial times (Sullivan
1990, Sullivan et al. 1990a), compared to 14% now.
These paleolimnological estimates suggest that currently acidic Adirondack lakes have decreased
a median of about 4 ueq l_' in (ANCG - Al.) for each kg ha'1 yr1 change in sulfur deposition. This
is slightly more than one-half of the median historical rate of change projected by MAGIC (7 /zeq
L'1 of ANC for each kg ha' yr"1) for acidic Adirondack lakes (Sullivan et al. 1992).

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A comparison between data coliacted in 1975 (Schofield 1976) and the ALSC survey data
collected in 1984-87 (Kretser et al. 1989) showed decreases in S042" (estimated from the charge
balance), NO,, and the base cation sum (C0), but no consistent pattern of change in ANC or Al
(Sullivan et al. 1992). Similarly, results from the 1000 lakes survey in Norway (Henriksen et al.
1988) suggested comparable decreases in CB and CA, presumably in response to lower S
deposition, leaving ANC generally unchanged. Comparisons for pH change in the Adirondack
lakes were complicated by partial air equilibration in the 1975 survey. Very low-pH lakes (< 5)
are not significantly influenced by partial pressures of C02 above atmospheric values, however,
and the data for these lakes did not suggest substantial changes in pH between the surveys
(Sullivan et al. 1992).
Decreases in [S042' + N03] (25 fieq L"1) were approximately equal to decreases in CB
(31 fieq L"'), suggesting that the predominant response of the lakes to decreased deposition and
sulfate concentrations had been decreased base cation mobilization (Sullivan et al. 1992).
Substantial changes in pH and ANC did not occur.
There are largely unquantifiable uncertainties related to seasonality, methodology, and
natural variability for this comparison. The data are internally consistent, however, and suggest
changes in lakewater [S042] that are similar to the estimated percent changes in sulfur deposition
(Husar et al. 1991). The data further suggest that substantial changes in acid-base chemistry
have not occurred despite an approximate 15-20% decrease in sulfur deposition during the
intervening period, in agreement with MAGIC estimates of future change. Neither has declining
sulfate concentrations in streamwater resulted in ANC-recovery in streams in the Catskill
Mountains. Stoddard and Murdoch (1991) found that in the streams lowest in ANC, a
combination of increasing N03" and decreasing (Ca2+ + Mg2+) concentrations appear to outweigh
the importance of declining S042' concentrations, and may be driving ANC downward.
Apparently, a larger percentage of deposited nitrogen is being passed through Catskill
watersheds which previously retained N strongly (Stoddard and Murdoch 1991).

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In comparing estimates derived from different approaches of past and future changes in
Adirondack lakewater chemistry in response to acidic deposition, it is important to consider
several factors (Sullivan et al. 1992):
•	Chemical changes estimated from paleolimnology or monitoring/survey data incorporate
all influences on the acid-base chemistry of lakewater, including land use. disturbances,
and climatic differences. Model estimates include only postulated or estimated changes in
acidic deposition as having influenced the lakewater chemistry.
•	The use of MAGIC, or any process-based model, for hindcasting requires assumptions
regarding historical deposition of all major ions. In addition to uncertainties regarding
historical sulfur deposition levels, base cation deposition has also likely changed by an
unknown amount, and the degree to which sulfur and base cation deposition are coupled
is unclear (Driscoll et al. 1989b; Chen and Gomez 1989).
•	Organic acids may have exerted a greater influence on lakewater pH during pre-industrial
times than they do currently because DOC and organic acid anion concentrations may
have decreased in response to increased organic acid protonation and increased
concentrations of Al (Aimer et al. 1974, Krug and Frink 1983, Davis et al. 1985, Kingston
and Birks 1990). In addition, the lower ionic strength during pre-industrial times likely
resulted in a larger pH effect from a given quantity and quality of organic acids.
•	Data sets are often not directly comparable because of differences in ANC definition or pH
measurement. For example, the defined ANC (CB - CA) used by MAGIC differs from
titrated ANC (ANCG) often used to calibrate paleolimnological transfer functions and
reported in surveys. The differences are due to the partially counteracting influences of Al
and organic acids on ANCG and their omission from (CB - CA). These differences can be
appreciable for acidic and low-ANC waters (Sullivan et al. 1989). Use of air equilibrated
versus field pH can cause large differences in measured pH values for waters having pH
> about 5 because of C02 effects.
Thus, there are uncertainties and limitations with all available approaches and inconsistencies
between data sets that often complicate direct comparisons between different methods.
Nevertheless, it is only through such comparisons that model output can be evaluated.
Traditional uncertainty analyses provide only an indication of model calibration uncertainty and do
not reflect uncertainty associated with the fundamental model assumptions and structures.
Furthermore, although comparison between models can yield very useful information, the
process-based models share a number of common assumptions (Reuss et al. 1986).
Natural variability and inconsistencies between methods are often larger than the chemical
changes under investigation. Although the changes may be biologically significant, detection and

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quantification of change is difficult. Under such conditions, it is particularly important to utilize a
variety of assessment tools that are not constrained by common assumptions.
Based upon intercomparison among all of the estimates available for past acidification of
Adirondack lakes, it appears likely that MAGIC model estimates of future chemical change
represent an upper bound for lakewater response. Furthermore, the MAGIC estimates of future
change are likely most realistic for the acidic and near-acidic systems. Lakes having higher ANC
may be generally less responsive to changes in acidic deposition than would be suggested by
the MAGIC model (Sullivan et al. 1992). Thus, the MAGIC model appears to be an appropriate
tool for generating quantitative estimates of acid deposition standards for the most sensitive
watersheds. In most cases, however, MAGIC may provide an overestimate of the responsiveness
of watershed systems to changes in sulfur deposition.

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V. ASSESSMENT OF SULFUR AND NITROGEN DEPOSITION LEVELS
REQUIRED TO PROTECT AQUATIC RESOURCES
A. REGIONAL DOSE/RESPONSE ASSESSMENT
1. West
The Sierra Nevada are particularly sensitive to potential acidic deposition effects because of
the predominance of granitic bedrock, thin acidic soils, large amounts of precipitation, coniferous
vegetation, and extremely dilute waters (McColl 1981; Melack et al. 1982, 1985; Melack and
Stoddard 1991). Similarly, Cascade and Rocky Mountain lakes are highly sensitive to potential
acidic deposition effects (Nelson 1991; Turk and Spahr 1991). There are no data to suggest that
lakes in these areas have experienced chronic acidification to date, and based on examination of
current chemistry it appears that chronic acidification has not occurred to any significant degree.
It is possible, however, that episodic effects have occurred under current deposition regimes.
Unfortunately, the data that would be needed for such a determination have not been collected to
a sufficient degree in acid-sensitive areas of the West to permit any regional assessment of
episodic acidification.
Episodes of surface water acidification involving increases in N03" concentration have been
reported for a few sites in the western United States (e.g., Loranger et al. 1986, Loranger and
Brakke 1988, Melack and Stoddard 1991). Maximum N03" concentrations in western lakes and
streams reported during episodes tend to be low, however, typically less than 15 //eq L"'
(Stoddard 1994). Although such episodic concentrations are quite low in comparison with many
sites in the eastern United States, increases in N03" concentration during episodes at Emerald
Lake in the Sierra Nevada have apparently been sufficiently high, when coupled with natural base
cation dilution, to drive lakewater ANC to zero on two occasions (Williams and Melack 1991a,
1991b; Stoddard 1994).
Forest stand age is associated with nitrogen retention by the forest system. Watersheds
having older trees seem more likely to leach N to a.higher degree than forests having younger

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trees (e.g., Vitousek 1977, Elwood et al. 1991). In general, forests in the eastern United States
have been logged on many more occasions and subjected to more intensive forest management
practices than forests in the western United States. In addition, stands of old growth trees, which
are generally absent from the eastern United States, can still be found scattered throughout areas
of the western United States. It is anticipated that, all other things being equal, the less
intensively managed western forests should be closer to a condition of nitrogen saturation than
are eastern forests which have been cut repeatedly and have been under more intensive forest
management for hundreds of years. Thus, results of studies conducted in the east regarding
thresholds of N saturation (e.g. Kahl et al. 1993) may underestimate the sensitivity of otherwise
generally comparable western forests to increased nitrogen loading.
The subsetting process for the five subpopulations in the West was successful in
consolidating most of the low-ANC lakes into more interpretable groups. However, several
smaller ranges were excluded from these summaries that are very important with respect to N
distributions. The Uintas, UT' and the Bighorn Mountains of central Wyoming have the greatest
percentages of high N03' lakes in the West, with 19% of the lakes with N03 > 10 fieq L"\ This is
a remarkably high percentage of lakes with measurable N03' for fall samples and indicates that
nitrate deposition in these areas may have exceeded the capability of these systems to assimilate
N. It is unknown if these concentrations of N03 represent impacts from anthropogenic sources or
if this constitutes an unusual natural condition associated with inhibited nitrate assimilation in
extremely cold alpine environments.
Of the regions under investigation for this study, the West is most susceptible to potential
acidification from acidic deposition. Because of the paucity of dose-response data for the region,
it is unclear what level of deposition would be appropriate for the protection of aquatic resources
from adverse effects. Based upon the weight of evidence, it is our professional judgement that an
appropriate standard for sulfur deposition would be less than 10 kg S ha'1 yr1 to protect against
chronic acidification in large areas of the west. A standard sufficient to protect against episodic

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acidification may be much lower than that, perhaps in the range of 5 kg S ha"1 yr'1. Furthermore,
in the most sensitive portions of the West (e.g., Sierra Nevada and Cascade Mountains), an
appropriate standard for protecting the most sensitive aquatic resources against chronic and
episodic acidification is probably below 5 kg S ha1 yr"1. Such estimates are highly subjective,
however, and should be considered as "best guesses" at this time. As noted earlier, use of a
deposition standard in the West is problematic because of the wide range in precipitation volume.
In some areas of the Pacific Northwest, precipitation volumes are so great (up to 4 m) that all but
the most conservative standards will be violated under natural conditions. Some consideration
needs to be given to an annual-weighted pollutant concentration standard or a combination
deposition/concentration standard rather than a stand-alone deposition standard in many areas of
the west.
Selection of appropriate standards for N is even more difficult than for S for the protection of
western lakes. In addition, it is likely that N deposition represents a more immediate threat to the
ecological health of high elevation western lakes than does S deposition. In view of the almost
complete lack of appropriate (i.e., seasonal) chemical data with which to evaluate nitrogen
acidification issues in the West, the most we can say at this point is that we do not know enough
to offer even "best guess" professional judgements regarding the selection of numerical standards
for N deposition.
2. Upper Midwest
Acidic lakes in the Upper Midwest are primarily small, shallow, seepage lakes that have low
concentrations of base cations, low [Al] and moderate [S042]. Organic anions, estimated by both
the Oliver et al. (1983) method and the anion deficit, were less than half the measured S042'
concentrations (Eilers et al. 1988a).
Groundwater flow-through lakes in the Upper Midwest were defined by Baker et al. (1991a)
as those having Si02 > 1 mg L'V Classification of seepage lakes into groundwater flow-through

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and recharge categories on the basis of [Si02] has been shown to result in a very clear distinction
of upper midwestern subpopulations and marked differences in [CB] and ANC between these
subpopulations. Groundwater flow-through lakes in the Upper Midwest were generally high
pH/ANC systems, largely due to substantial groundwater inputs of base cations (e.g., Baker et al.
1991a). Only 6% of these lakes had ANC <_ 50 peq L'1 and none were acidic. They are therefore
of relatively little interest with respect to potential past changes in acid-base status in response to
acidic deposition. Groundwater recharge lakes in the Upper Midwest were more common (71%
of the seepage lakes), and were more frequently low pH/ANC. Five percent were acidic and 9%
had pH < 5.5. Nearly 90% of upper midwestern lakes that had ANC <. 50 /zeq L"1 were in
this hydrological category (see e.g., Baker et al. 1991a). The modest influence of watershed
processes on the chemistry of groundwater recharge lakes results in minimal weathering
contributions, as reflected by the low concentrations of Si02 and Ca2+ (Schnoor et al. 1986).
Although paleolimnological data suggest that some upper midwestern lakes have acidified
since pre-industrial time, there is little paleolimnological evidence indicating substantial
widespread acidification in this region (Kingston et al. 1990; Cook et al. 1990). It is likely that land
use changes and other human perturbations of upper midwestern lakes and their watersheds
have exerted more influence on the acid-base chemistry of lakes than has acidic deposition
(Eilers et al. 1989, Kingston et al. 1990, Sullivan 1990). This result is not unexpected because
acidic deposition has been much smaller in magnitude in the Upper Midwest than in most areas
of the eastern United States (Husar et al. 1991). It is clear, however, that the portion of the region
most likely to have experienced acidification from acidic deposition is the Upper Peninsula of
Michigan, where acidic seepage lakes are particularly numerous (Baker et al. 1990a); acidic
deposition is highest for the region, and the [S042']/[CB] ratio is commonly > 1.0 (Figure IV.B.2).
The percentage of acidic lakes in the eastern portion of the Upper Peninsula of Michigan (east of
longitude 87°) is 18% to 19% (Schnoor et al. 1986; Eilers et al. 1988a).

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Paleolimnological data are available for only two lakes in upper Michigan, McNearney and
Andrus Lakes. McNearney Lake may have been naturally acidic and appears to be atypical for
the region. Andrus Lake is inferred to have experienced declines in pH and DOC since pre-
industrial times that could be related to acidic deposition (Kingston et al. 1990). It is likely that
other lakes in this subregion have also experienced recent acidification. Data are lacking,
however, with which to quantify the amount of acidification that may have occurred in the past or
the dose/response relationships of these systems. In addition to the scarcity of paleolimnological
data within the portion of the Upper Midwest most likely to have experienced widespread
historical acidification, there is also a paucity of basic biogeochemical data on the response of the
major lake type in this region (seepage) to atmospheric inputs of sulfur and nitrogen.
The spatial distributions of lakewater chemical variables across a longitudinal gradient in the
Upper Midwest for low ANC subsets of drainage lakes and groundwater recharge seepage lakes
(Figures IV/.B.1, IV.B.2) constitute perhaps the best evidence available that many of the most
sensitive lakes in the eastern portion of this region have acidified. In the absence of additional
paleolimnological data for these systems of most interest, however, it is difficult to substantiate in
terms of magnitude much regional acidification in the Upper Midwest.
Nitrogen deposition does not appear to be an important issue for sensitive aquatic resources
in the Upper Midwest. This is likely attributable to the fact that snowmelt is less important to the
acid-base chemistry of sensitive (i.e., seepage) lakes in this region; sensitive aquatic resources
are not found at high elevations, where low summer temperatures may limit biological activity;
and hydrologic retention times are long. Sulfur deposition appears to be of greater importance in
this region, and potential chronic effects are of greater interest than episodic effects because of
the nature of the hydrology of sensitive resources in the region. Our best professional judgement
for establishing a preliminary standard for sulfur deposition in the Upper Midwest is that current
deposition in the eastern portion of the region (~ 5 kg S ha"' yr'1) is a reasonable approximation
of the deposition level required to protect those most sensitive aquatic receptors. Resources in
March, 1994
Page 112

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the western portion of the region are less sensitive, however, and an appropriate standard for S
deposition would be much higher. Because deposition in this region has been decreasing in
recent years, we believe that acidic deposition is not an important environmental concern in the
Upper Midwest at this time.
A sulfur deposition standard has been in effect in Minnesota since 1986. The Minnesota
standard was based on the Acid Deposition Control Act, passed by the state legislature in 1982,
which required the Minnesota Pollution Control Agency (MPCA) to identify natural resources
within the state that were threatened by acid deposition and to develop both an acid deposition
standard and an emissions control plan. Small, poorly-buffered lakes in northcentral and
northeastern Minnesota were identified as the resources at greatest risk. Based on model
simulations, MPCA selected a threshold pH for precipitation of 4.7, below which damage to
aquatic biota was thought to occur with prolonged exposure. This threshold pH was correlated
with S042' deposition data, and a standard was determined that allowed no more than 11 kg ha'1
of wet S042' to be deposited during any 52-week period (3.7 kg S ha'1 yr'1) (MPCA 1985). This
standard is fairly stringent. In fact, six of twelve monitoring sites in Minnesota exceeded the
standard in 1992 (Orr 1993).
3. Florida
Differences in major ion chemistry of lake waters occur across Florida, and are related
primarily to differences in precipitation and evaporation (Pollman and Canfield 1991). The most
dilute lakes occur in the Florida Panhandle where precipitation is highest (ca. 150 cm/yr).
The most acidic of the paleolimnological study lakes (Lakes Barco and McCloud) are found
in or just south of the Trail Ridge in northcentral Florida. The chemistry among the northcentral
Florida lakes is somewhat more variable than for lakes in the Panhandle because of surficial
geologic differences and variable contributions from wetlands occurring over small distances. For
example, Suggs Lake, which is less than 2 km from Lake Barco, is highly colored (28 mg L"'

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DOC) with extremely high CI" and Na+ concentrations, and comparatively low S042"
concentrations. Lake Barco, which lies upland of Suggs Lake, is a more dilute, Clearwater lake
yet has much higher S042' concentration and is more acidic (pH 4.48).
Historical changes in Florida lakewater chemistry, as inferred from diatoms, showed a distinct
geographical pattern. All five of the paleolimnological study lakes in the Trail Ridge region
showed some evidence of acidification, some strongly linked in timing to both the period of
increasing acidic deposition and increased water consumption. Trail Ridge lakes showed diatom-
inferred ApH ranging from -0.2 (McCloud) to -0.9 (Suggs; Table IV.C.2). No clear evidence of
acidification was observed for lakes in Ocala National Forest (three lakes) or the Panhandle (eight
lakes), except Lake Five-O, where gross hydrological change was implicated. It is most likely that
several factors have caused the recent acidification of lakes in the Trail Ridge area suggested by
the diatom data. Acidic deposition is implicated, but changing lake stage and the linked
phenomenon of evapoconcentration may also be important (Sweets et al. 1990).
Diatom-inferred historical changes in pH for all lakes'in the Florida Panhandle, except Lake
Five-O, were less than -0.10 units. These results appear surprising insofar as the Panhandle
seepage lakes are the most dilute lakes in Florida, and have been believed to receive minimal
hydrologic in-seepage (ca. 1-3 percent of total hydrologic budget; cf. Baker et al. 1988). Recent
groundwater monitoring data collected adjacent to Lake Five-O suggest, however, that
groundwater may contribute one-third to one-half of the overall hydrologic budget of this lake
(Pollman et al. 1991). Calibrated inflows based on CI" balances for Panhandle lakes also
suggested substantial groundwater inflows, ranging from 10 (Moore Lake) to 29 (Lofton Ponds)
percent (Pollman and Sweets 1990).
Superimposed on the complex heterogeneity of Florida lakes is a high incidence of
anthropogenic disturbance. Of the 159 total lakes sampled by ELS-I in Florida, all but 37 were
judged by Baker et al. (1988) to have substantial shoreline or watershed disturbances, mostly
related to agriculture. Besides the increased atmospheric deposition in Florida in the 1950s, other

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changes have also occurred. The human population has increased markedly, as has freshwater
withdrawal from the Floridan aquifer (Fernald and Patton 1984; Aucott 1988). As a result, the
potentiometric head has declined substantially in the Trail Ridge area (Healy 1975; Aucott 1988;
Pollman and Canfield 1991).
For undeveloped lakes in the northcentral peninsula, lakewater chemistry is consistent with
an hypothesis of acidification by acidic deposition. (Hendry and Brezonik 1984; Eilers et al. 1988b;
Baker et al. 1986; 1988). Evaporative concentration of modest amounts of acidic deposition, and
in-lake retention of sulfate and nitrate appear to be important processes. However, Eilers et al.
(1988b) concluded it is unlikely that the mechanisms of acidification of Clearwater lakes in Florida
and the linkages to atmospheric deposition will be satisfactorily understood until the hydrologic
pathways are better known. Slight differences in groundwater inputs can have a major influence
on base cation supply and lakewater chemistry in these precipitation-dominated seepage
systems. Based on limited paleolimnological data, it appears that recent acidification of lakes in
Florida may have been restricted to the Trail Ridge district. Furthermore, it is unclear to what
extent recent acidification of lakes in the i rail Ridge district may be attributable to acidic
deposition, as compared to other anthropogenic activities, especially groundwater withdrawal.
Regarding recommendations for selection of appropriate standards for the protection of
aquatic resources in Florida from adverse effects of acidic deposition, we can offer little guidance.
We have seen no evidence that current levels of N deposition exceed watershed capacities to
assimilate nitrogen. Evidence suggestive of possible acidification from S deposition is confined to
a small area in the Trail Ridge region of northcentral Florida, which receives relatively high levels
of sulfur deposition (Figure IV.A.1). Recent acidification of Trail Ridge lakes could also be
attributed, in part or in whole, to groundwater withdrawals for consumptive use.
It would not be expected that a quantitative deposition standard for the Upper Midwest or
Florida would necessarily be similar to an appropriate standard for the Adirondacks. Seepage
lakes predominate in the Upper Midwest and Florida, and are not expected to respond in a similar

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fashion to acidic deposition as have the predominantly drainage systems of the Adirondacks. As
a consequence of this basic hydrological difference, concentrations of dissolved Al are
considerably lower in seepage lakes, as compared to drainage lakes of comparable pH (Sullivan .
1991, Munson and Gherini 1991). The fisheries responses are also different. Acid-tolerant
species prevail in Florida and the Upper Midwest (Pollman and Canfield 1991, Cook and Yager
1991), whereas acid-sensitive fish species are more common in the Adirondacks (Baker and
-Christensen 1991).
4. Eastern Canada
Data collected from lakes in the vicinity of the smelting operations near Sudbury, Ontario,
have provided a wealth of information with which to quantify both acidification and recovery
responses. Particularly useful data have included the results from both paleolimnological and
monitoring/survey studies. Documented recent losses of fish populations in eastern Canada,
attributable to acidic deposition, have generally been restricted to the Sudbury region (Kelso et al.
1990, Matuszek et al. 1992, Conlon et al. 1992). In addition, research in the Sudbury area has
provided some of the best data available with which to evaluate the reliability of diatom inferences
of pH change. However, interpretation of dose-response data from Sudbury is complicated by
the very high levels of acidic deposition received by many of the watersheds that have been
studied. The observed changes in deposition and associated effects, have been substantially
greater than would be experienced during any realistic future deposition scenarios for the regions
under investigation for this report. It is therefore difficult to use such data in the analyses
presented here.
Beyond the general vicinity of Sudbury, quantitative data on acidification response are scarce
for eastern Canada (Cook et al. 1989). Although current chemistry has been investigated to a
considerable degree through lake surveys, the resulting data are insufficient for quantitative dose-
response assessment.

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In 1983, the Canadian Council of Resource and Environment Ministers established a target
loading of 20 kg ha1 yr'1 of wet SO„2' deposition for sensitive areas of eastern Canada. This
target was based on the review of US-Canada (1983) and a desire to maintain surface water pH
above 5.3 on an annual basis. It corresponds approximately to 8 kg ha'1 yr1 of total S deposition,
assuming dry deposition equal to 15% of total sulfur deposition (e.g., RMCC 1990). Subsequent
analyses by RMCC (1986), using the ESSA-DFO model of Marmorek et al. (1985), predicted that
further deposition reduction would' be required to 12 kg ha"1 yr"1 wet S042' deposition to prevent
additional acidification of lakes to pH below 5.0. RMCC (1990) re-examined the critical and target
loads issues and concluded that the adoption of a single critical load for SO„2 deposition in
eastern Canada was not appropriate. They determined that critical loads were site specific, due
to variability in watershed and climatic characteristics. Based on a review of studies on biological
effects, the threshold pH of 5.3 was re-evaluated. The threshold value of pH = 5.3 had been
adopted for the protection of sport fish species, whereas aquatic ecosystem structure and
function were believed to be affected at higher pH values. RMCC (1990) revised the threshold pH
to 6.0 as more appropriate for the protection of fish and other aquatic biota, based on the weight
of existing evidence on effects. Selection of pH = 6.0 as a threshold was somewhat problematic,
however, because many lakes were estimated to have had pH less than 6.0 in the absence of
acidic deposition. These were excluded by RMCC (1990) from critical load determinations.
Critical loads were assigned on a regional basis, stratified into 22 aggregations of watersheds
(Figure V.A.1). The aggregations (subregions) were selected to minimize within-group variance,
and maximize between-group variance, in lakewater conductivity. Geological characteristics and
sulfur deposition were also used in defining boundaries. Steady state models were used to
estimate the critical load for each aggregate, and to predict the percentage of lakes in each
aggregate that would be acidic under a deposition equal to the previously recommended target
load of 20 kg ha'1 yr1 (Table V.A.1). In about one-third of the watershed aggregates, more than
5% of the lakes are currently acidic (Table V.A.1). RMCC (1990) estimated that in four-fifths of the

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E&S Environmental Chemistry, Inc.
Critical load of total sulfur
deposition (kg S ha' yr')
C^l
- 3-1
rm
> 3,1 - 4.6
ZZ2
> 4.6 - 6.1
a
> 6.1 - 7.7
VTTTTX
> 7.7
Figure V.A.1
Critical load of total sulfur deposition for watershed aggregates of southeastern
Canada, as determined by the Canadian 1990 Assessment (modified from RMCC
1990).

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22 watershed aggregates, more than 5% of the lakes would be acidic under a loading rate of 20
kg ha'1 yr"1 wet sulfate deposition (~ 8 kg ha1 yr' total S).
Estimates of critical loads were derived from a combination of six steady state water
chemistry models, each based on the model described by Marmorek et al. (1985) or Schnoor et
al. (1986). Neither the ESSA-DFO model (Marmorek et al. 1985) nor the TD model (Schnoor et al.
1986) included an organic acid representation. These models were therefore linked with the
organic acid models described by Oliver et al. (1983) and Lam et al. (1989). The resulting six
possible combinations of steady state ANC models with or without linkage to an organic acid
Table V.A.1.
Percentage of sampled lakes within subregions (termed tertiary watershed

aggregates) in southeastern Canada at or below specific ANC values.
Percentages

include both those observed under current deposition and those predicted using:

steady-state models for 20 kg ha'1 yr'1 wet S042'
deposition. Only lakes having

sufficient data to permit model application have been included. The twenty-two

watershed aggregates were delineated by RMCC (1990) and are shown in Figure

V.A.1.
(Source: RMCC 1990)





Percent Observed

Percent Predicted
Watershed

(current deposition)
(20 ka ha'1 vr:' wet SO.''
deposition)*
Aggregate
n
ANC < 0 ANC <. 50
ANC < 0 ANC < 50
1
280
14
83
49
84
2
59
10
80
59
85
3
89
1
43
32
55
4
33
36
85
73
88
6
60
0
32
52
95
7
17
6
77
71
100
8
. 30
7
47
30
70
10
103
1
61
69
96
11
10
0
30
30
40
12
64
0
3
3
8
13
251
2
35
16
57
14
408
0
54
7
46
15
529
1
29
1
14
16
80
0
23
5
24
17
1009
5
47
4
16
19
287
33
72
37
66
20
541
0
1
2
4
21
36
0
0
28
47
22
16
6
44
13
69
• Approximately eauals 7 7 kg S ha'1 vr"' of total S deDOsition

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model were employed for model
estimates. The most appropriate
model combination for a given lake
was selected according to selection
rules using the RAISON expert
systems (Lam et al. 1988, 1989a) and
including organic acids if color > 30
Hazen units or DOC > 4 mg L'\ ANC
values computed by application of the
expert system and steady state
models were converted to pH using an
inverse hyperbolic sine function
according to Small and Sutton (1988).
The critical load of sulfur
deposition required to achieve pH less
than or equal to 6.0 in 5% of the
sampled lakes of each watershed
aggregate was estimated using these
modeling approaches. Critical loads
less than about 3 kg S ha"' yr'1 were
specified for almost half of the watershed aggregates (Table V.A.2, Figure V.A.1).
B. RECOMMENDATIONS
1. Quantification of Chronic Acidification Responses
Measured changes in surface water chemistry in areas that have experienced short-term
(< 20 yr) changes in chemical constituents in response to changes in mineral acid inputs are
Table V.A.2.
Estimated total sulfur deposition

required to achieve pH less than or

equal to 6.0 in 5% of sampled lakes in

watershed aggregates of eastern

Canada (Source: RMCC 1990/

modified to include estimated dry

deposition);


Percent of Lakes
Watershed

Having Historical Critical Load
Aggregate.
n
pH < 6* of Sulfur0
1
280
52 < 3
2
59
22 <4
3
89
7 <3
4
33
42 <3
6 ¦
60
12 <3
7
17
0 <3
8
30
3 <3
10
103
13 <3
11
10
0 <3
12
64
0 <5
13
251
1 <3
14
408
2 4
15
529
0 >8
16
80
0 5
17
1009
0 6
19
287
0 <6
20
541
0 >8
21
36
6 <3
22
16
6
" These lakes were excluded from the critical loads
estimates


b Critical loads expressed as kg of total S ha1 yr
assuming dry deposition equal to 15% of wet in all
subregions except 19 (100%) and 2 (35%).

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available from a number of sources. These include results from manipulation experiments and
changes in acidic deposition. Proportional changes in ANC (expressed as [HC03' - H*]), [CB],
and [AIJ, relative to changes in [SO/] or [SO/ + N03], are summarized in Table V.B.1 for lakes
and streams in which such changes have been measured. They include lakes in the Sudbury
region of Ontario, the Galloway lakes area of Scotland, a stream site at Hubbard Brook, New
Hampshire, and catchment manipulation experiments in the RAIN project in Norway and Little
Rock Lake in Wisconsin. Most of the observed changes are coincident with decreased acidic
deposition, and it is unclear whether acidification and recovery are symmetrical. F-factors in the
range of 0.5 to 0.9 are apparently typical for lakes having low [CB], although lower values (0.35 to
0.39) were observed for the highly sensitive catchments at Sogndal, Norway, which are
characterized by thin soils and much exposed bedrock, as is common in many areas of southern
Norway. The proportional change in ANC relative to [S042' + N03 ] change was variable, within
the range of 0.1 to 0.5. The proportional change in Al was smaller, ranging up to 0.15.
Relatively early in the international efforts to quantify the acidification response, Henriksen
(1982) proposed that F-factors for softwater lakes would be in the range 0 to 0.4. More recent
research (e.g., Table V.B.1) has shown this earlier estimate to be too low in most cases. Based
on measured values, only the most sensitive systems, for example at Sogndal, exhibit F-factors
below 0.4. Nevertheless, a number of investigators continue to state that changes in base cation
concentrations will only account for up to about 40% of the change in S042' concentration (e.g.,
Webster et al. 1993).
In addition to the measured acidification and recovery data presented in Table V.B.1, there
are several other sources of quantitative or semiquantitative data with which to evaluate the
general applicability of the measured results that are available. These include the results of
space-for-time substitution (Table V.B.2), diatom-inferences of historical acidification (Table V.B.3),
and results of process-based model hindcasts or future forecasts (Table V.B.4). Each of these
methods has its own assumptions and limitations, and none are as robust as results of actual

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Table V.B.i. Measured short-term changes in surface water chemistry associated with changes in mineral acid anion concentrations.
(Units in/ieq L ')




Initial

A (HCOj'-H*)
ac;
A Al

Site
Location
Period
Type
pH
A SO,2
A SO/
A S043
A S04a
Ref
Clearwater Lake
Sudbury. Canada
1973-77 to
1984
Recovery
4.2
-175
019
0.66"
0.15
1
Swan Lake
Sudbury. Canada
1977 to 1982
Recovery
4.0
-360
0 26
0.67
0.07
1
Baby Lake
Sudbury. Canada
1968-72 to
1903
Recovery
4.05
-750
0.12


2
Whitepine Lake
Sudbury, Canada
1980 to 1908
Recovery
5 4
-42
0 24

0.05
6
Laundrie Lake
Sudbury. Canada
1974-76 to
1979-03
Recovery
4.7
-50
0 24


3
Florence Lake
Sudbury, Canada
1974-76 to
1979 03
Recovery
4.6
-42
0 22


3
Average of 37
lakes having
plH < 5.5
Sudbury. Canada
1974-76 to
1979-03
Recovery
4.7
-42
0.15


3
Average ol 105
lake trout lakes
Sudbury, Canada
1980 to 1907
Recovery

-45
0 51


6.11
Average ot 50
lakes
Galloway,
Scotland
1979 to 1908
Recovery
5.4 ±
0.71
-76*
0 13
0.84
~0.06
4.8
Little Rock Lake*
Wisconsin
1983 to 1909
Acid Addition
6^6
94
0.44
0.53

10
S0G2 catchment
Sogndal, Norway
1904 to 1987
Acid Addition
5.5
28*
0.46
0.39
0.11
5
SOG4 catchment
Sogndal, Norway
1984-1987
Acid Addition
6.0
20*
0.35
0.35
0.15
5
KIM catchment
Risdalsheia.
Norway
1904 to 1907
Acid Exclusion
4.1
-139"
0.09
0.55
0.05
5,7
Bear Brook
Maine
1907 to 1992
Acid Addition
5.6
62*
0.14
0.51
0.20
12
Hubbard Brook
New Hampshire
1969 to 1979
Recovery
4.0
-30'
0.15
0.91
-
9
*	also includes NOa
" A CJ A SO' calculated by difference, assuming that the proportional changes in alkalinity, CB. and Al sum to 1.0.
c Changes in the organic anion contribution to acidity were important at this site, where DOC was very high (~ 1250 pM).
" 1 - Dillon et al. 19B6; 2 - Hutchinson and Havas 1966; 3 - Keller et al. 1906; 4 - Wright 1908b; 5 - Wright et al. 1900b; 6 - Gunn and Keller 1990;
7 - Wright et al. 1909; 8 - Wright et al. in review, 9 - Sullivan 1990; 10 - Sampson et al., in press; 11 - J. Gunn, personal communication; 12 - Norton et al. 1993
*	Little Rock Lake experiment involved manipulation of lake only
m >
fiO (A
« «y jg
< o 3
i if ffl
O ® 3
3 " -
3 > a
3 c o
ET B ®
— seTJ
O ° 2
3 8 §
a 5 •"
^2 1
-Si
3 (0
P o
(/>
c
q-
o
<0
 *

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lahle V.b.L'.
Interred long term regional changes in suitace water chemistry associated with estimated changes in mineral acid
anion concentrations, using the technique of space-for-time substitution.

AANC
A Cfl
A Al


Region
ASOl
A SO<~
A SO7<
Reference
Comments
NE U.S.
0.13
0.54
0.07
Sullivan et al. 1990
Analysis restricted to lakes
having current ANC £
25 fteq L'
S. Norway
0.22


Brown & Sadler 1981
Regional data set (n=471) .
S. Norway

0.82

Wright 1988,
Sullivan 1990
Lakes located across
depositional gradient from Bykle
to Mandal
m

>
CP

(A
(/>
O
$
m
3
"O
(0
§
<
o

o

i)

3
c
o
2
G»
c.
©
¦o
o
rr

c
c
3
a
o

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Table V.B.3. Diatom-inferred long-term changes in lakewater ANC as a fraction of
estimated historic changes in lakewater sulfate concentration.
Number of
Region of Lakes
A ANC
A S042"
Reference
Comments
Adirondacks, NY 48
sampling
0.11
Sullivan et al. 1990
Statistical
Adirondacks, NY 25
0.18
Sullivan et al. 1990
Acidic lakes only1
Florida (L. Barco, 2
Suggs)
0.27
This report
Seepage lakes
1 The set of 25 acidic lakes was part of the regional data set of 48 lakes presumed to be
acid-sensitive
field measurements of response. A major advantage of these alternative sources of quantitative
data, however, is that they primarily reflect acidification, rather than recovery, scenarios.
Results of space-for-time substitution must be interpreted with caution. This approach is
based on the assumptions that changes in chemistry across space, for example from low to high
levels of acidic deposition, reflect changes that occurred over time as deposition increased from
low to high. It is implicitly assumed that the waters included in the analysis were initially
homogeneous in their chemistry, and also that potentially important factors other than deposition
(e.g., soil characteristics, land use impacts) do not co-vary with deposition. Results should
therefore be considered only semi-quantitative. Nevertheless, available data using this method
(Table V.B.2) appear very similar to results of measured values shown in Table V.B.1.
Diatom-inferences of change in ANC from pre-industrial times to the present have been
reported for a regional population of Adirondack lakes (Sullivan et al. 1990a), and for two lakes in
Florida that have shown clear decreases in pH and ANC in recent decades (Sweets 1992).
Proportional changes in diatom-inferred ANC as a fraction of assumed increases in S042'
March, 1994
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Table V.B.4. Dynamic model (MAGIC) estimates of F-factors (or hindcast or future forecast projections of acidification or
recovery responses.
- Region
Type of
Simulation
Number
of Lakes
or Streams
F-Factor
Median 5th %ile
Reference
Adirondacks	Hindcast
Adirondacks	50 yr forecast, w/50%
reduction in S deposition
Wilderness lakes, Forecasted 3-fold increase
Western U.S.	in S deposition
Bear Brook, ME
Response to experimental
watershed acidification
33
33
15
0.56	0.25	Sullivan et al., in review
0.73	0.39	Sullivan, unpublished
0.34	0.03	Eilers et al. 1991
0.85	Norton et al. 1992

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concentration since pre-industrial times show estimates ranging from 0.1 to 0.3 (Table V.B.3), in
close agreement with measured values (Table V.B.1).
Dynamic model estimates ot F-factors for watersheds in the northeastern United States, using
the MAGIC model, show reasonably close agreement with measured F-factors for acid-sensitive
systems (Table V.B.1, V.B.4). Model generated median values of the F-factor ranges from 0.56 to
0.85, and values of the 5th percentile of Adirondack lake projections (0.25 to 0.39) were
reasonably comparable to the measured values at the highly sensitive Sogndal site (0.35 to 0.39).
MAGIC forecasts for western lakes, however, yielded estimated F-factors that were substantially
lower (median 0.34, 5th percentile 0.03; Table V.B.4). It is not clear how representative these
forecasts might be for western lakes in general, or how accurate the estimates are for the
modeled lakes. Nevertheless, these comparative data suggest that western systems are as
sensitive, or perhaps more sensitive, than any of the watersheds for which acidification and/or
recovery responses have been more rigorously quantified.
Diatom estimates of pH have been compared with measured pH values at numerous lake
sites where changes in acid-base status have occurred. Such validations of the diatom approach
have been performed for lakes that have been acidified and lakes that have recovered from
acidification or have been limed in Canada (e.g., Dixit et al. 1987, 1991, 1992a), Sweden (e.g.,
Renberg and Hultberg 1992), and Scotland (e.g., Allot et al. 1992). Diatom-inferred pH histories
generally agree reasonably well with the timing, trend, and magnitude of known acidification and
deacidification periods. In several cases, however, the sedimentary reconstructions were slightly
damped in comparison with measured values. That is, the diatom reconstructions did not fully
reflect the magnitude of either the water pH decline or subsequent recovery.
For example, Renberg and Hultberg (1992) compared diatom-inferred pH reconstructions
with the known pH history for several decades at Lake Lysevatten in southwestern Sweden. The
diatom-inferred pH history agreed well with both the acidification period of the 1960s and early
1970s and also the liming that occurred in 1974. The magnitude of pH change inferred from

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sedimentary reconstructions was slightly smaller, however, than the measured changes in pH for
both acidification and deacidification.
Allot et al. (1992) found diatom reconstructions of pH recovery in the deacidifying Round
Loch of Glenhead, Scotland to be somewhat smaller than the measured pH recovery since the
late 1970s. pH reconstructions from the sediment cores showed an average recovery of 0.05 pH
units. Measured increases in pH between 1978-79 and 1988-89 averaged 0.23 pH units. The
authors attributed this difference to attenuation of the reconstructed pH record due to sediment
mixing processes.
Dixit et al. (1992a) analyzed sedimentary diatoms and chrysophytes from Baby Lake
(Sudbury, Ontario) to assess trends in lakewater chemistry associated with the operation, and
closure in 1972, of the Coniston Smelter. Extremely high sulfur emissions caused the lake to
acidify from pH « 6.5 in 1940 to a low of 4.2 in 1975. Following closure of the smelter, lakewater
pH recovered to pre-industrial levels. The diatom-inferred acidification and subsequent recovery
of the lake corresponded with the pattern of measured values. However, the diatom-inferred pH
response was more compressed and did not fully express the amplitude of the pH decline or the
extent of subsequent recovery.
It is not known why diatom-inferences of pH change are often slightly attenuated relative to
measured acidification or deacidification. Possible explanations include the preference of many
diatom taxa for benthic habitats where pH changes may be buffered by chemical and biological
processes. Alternatively, such an attenuation could be a result of sediment mixing processes. It
is thus not surprising that MAGIC model simulations that included organic acid representations
estimated pre-industrial pH values slightly higher than diatom-inferred values for Adirondack lakes.
It is not possible to determine which method provides estimates closer to reality. It is reassuring,
however, that they provide results that are generally in reasonable agreement.

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a. Standards for Sulfur and Nitrogen
Consideration of the efficacy of adopting one or more acid deposition standards for the
protection of surface water quality from potential adverse effects of sulfur and nitrogen deposition
is a multifaceted problem. It requires that sulfur and nitrogen be treated separately as potentially-
acidifying agents, and that separate estimates for each be generated for all individual, well-
defined regions or subregions of interest. Appropriate criteria must be selected as being
indicative of damaged water quality, for example ANC or pH. Once a criterion has been selected,
a critical value must be estimated, below which the criterion should not be permitted to fall. For
example, if the selected criterion is surface water ANC, one could specify that ANC should not be
permitted to fall below 0, 20, or 50 fieq L1 in response to acidic deposition (e.g. Kamari et al.
1992). Selection of critical values for ANC or pH is confounded by the existence of lakes and
streams that are acidic or very low in pH or ANC due entirely to natural factors, irrespective of
acidic deposition (Sullivan 1990). In particular, low contributions of base cations in solution, due
to low weathering rates and/or minimal contact between drainage waters and mineral soils, and
high concentrations of organic acids contribute to naturally low pH and ANC in surface waters.
,Other factors also can be important in some cases, including the neutral salt effect (cation
retention) and watershed sources of sulfur.
Acid deposition standards might be selected on the basis of protecting aquatic systems from
chronic acidification; conversely episodic acidification might also be considered, and would be of
obvious importance in regions where hydrology is dominated by spring snowmelt. Thus,
selection of appropriate acid deposition standards involves consideration of a matrix of factors, as
outlined in Table V.B.5.
1) Sulfur
Sulfur deposition is a potential concern in all of the regions under investigation for this report.
Some degree of chronic acidification attributable to sulfur deposition has occurred in southeastern

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Table V.B.5. Factors that should be considered for selection of acid
deposition standards for the protection of surface water
quality.
Factors for Consideration
Possible Options
Acidifying agent
Nitrogen or sulfur
Regional delineation
Region- or subregion-specific standards
Temporal response
Chronic or episodic acidification
Damage criterion
ANC or pH
Critical values for criterion
ANC < 0, 20, 50, fieq L'1
pH < 5, 5.5, 6
Canada, in the eastern portion of the Upper Midwest region, and possibly in the Trail Ridge
region of north-central Florida. Regional quantification of the amount of acidification that has
occurred in southeastern Canada and the Upper Midwest is not possible with existing data.
Although more quantitative (paleolimnological) data are available for northcentral Florida, and
consequently historical changes in lakewater pH are better documented, the cause of recent
acidification in some Florida lakes cannot be definitively ascribed to acidic deposition. Substantial
groundwater withdrawals from local aquifers might explain part, or all, of the historical changes in
pH.
Although acidification and recovery responses have been well-documented and quantified in
the vicinity of large industrial point sources of atmospheric sulfur in the Sudbury area of Ontario,
data for regional assessment of acidification for Canada outside the Sudbury area are generally
lacking. However, detailed watershed studies in portions of southeastern Canada have yielded
results that are generally consistent with results of studies conducted in the Adirondack
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Mountains. There are a number of similarities that would suggest reasonable comparability
between these regions with respect to their sensitivity to acidic deposition, including:
•	widespread occurrence of granitic bedrock and shallow, acidic soils
•	abundance of dilute, low-pH drainage lakes
•	variable contributions of organic acidity to surface water acid-base status
•	importance of snowmelt to watershed hydrologic budgets
•	generally conservative behavior of S042' in watershed soils
•	importance of base cation dilution and nitrate enrichment as episodic processes
In the absence of additional regional data, therefore; we recommend that first-approximations of
acid-sensitivity for eastern Canada should be based on more quantitative results obtained in the
Adirondack Mountains. Model-based estimates of standards required to protect sensitive
Adirondack watersheds from adverse effects of sulfur deposition will be provided in a companion
ERL-C report being prepared by M.R. Church (personal communication). We suggest that
standards derived from these analyses of Adirondack watersheds be used as interim guidelines
for southeastern Canada.
In the Upper Midwest and Florida, seepage lakes constitute the potentially-sensitive
resources of interest. It is difficult to make direct comparisons of deposition and potential impacts
between these regions, however. Interpretation of deposition impacts is confounded by the
importance of natural marine deposition of S042" and CI" in Florida and also by the enhanced,
importance of evapoconcentration in Florida lakes, which increases the acidity of weakly acidic
solutions (e.g., Munson and Gherini 1991). It is likely that an appropriate sulfur deposition
standard for the Upper Peninsula of Michigan would be somewhat less than peak deposition
values recorded in the 1970's, although it is not possible to quantify how much less based on
available data. Furthermore, sulfur deposition in-this region has been declining steadily in recent
years, and will therefore likely be of less concern in the future than in many, other regions of the

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country. As an interim guideline, we suggest use of a standard for sulfur in the range of 5 kg S
ha'1 yr"1, which approximates current deposition in the eastern portion of the region.
Based on analysis of available sulfur dose-response data for sensitive watersheds worldwide
(Tables V.B.1 to V.B.4), it is clear that proportional changes in ANC and base cations in drainage
waters in response to changes in sulfur inputs are highly variable. Documented F-factors (ACe -r
A[S042"]) are generally above 0.5, although lower values have been found. Perhaps the best
available estimate of an appropriate F-factor for highly sensitive watersheds, such as are found
throughout the western United States, would be based on the experimental values obtained at
Sogndal, in western Norway (near 0.4). This alpine watershed exhibits substantial areas of
exposed bedrock, and contains shallow acidic soils. As such, it appears to be a reasonable
surrogate for sensitive watersheds in the West. Although MAGIC model projections for western
lakes (e.g., Eilers et al. 1991) suggest that some watersheds may exhibit values for the F-factor
lower than 0.4, assessments using multiple approaches have concluded that MAGIC projections
may represent upper bounds for watershed acidification response (NAPAP 1991, Sullivan et al.
1992). We therefore recommend a value for F of 0.4 as most likely representative for highly
sensitive aquatic systems in the western United States. As a worst case scenario, a value as low
as perhaps 0.2 may not be unreasonable for extreme cases of acid sensitivity. Assuming such a
high level of sensitivity (F = 0.2) would certainly not be appropriate for watersheds in the
northeastern United States, based on all available information. It must be recognized, however,
that surface waters in the western United States probably are among the most sensitive in the
world to inputs of acidic deposition (Eilers et al. 1990, Melack and Stoddard 1991).
The first and fifth percentiles of measured ANC for the subregions of the West under
investigation are presented in Table V.B.6. Assuming F = 0.4, lakewater SO„2' concentration
would only have to increase by a modest amount, to drive the most sensitive lake to zero ANC in
each of the subregions (Table V.B.6).

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Table V.B.6. First and fifth percentiles of the regional lake ANC distributions for
subregions of interest in the western United States, and estimates of. the
increase in lakewater S042' concentration that would be required to drive
chronic ANC to zero. Units are in ^eq L'\


Current Lake
ANC
A S042"
to drive ANC to 01
Subregion
1st
5th
1st
5th
Sierra Nevada
15
16
25
27
Cascade Mountains
11
18
18
30
Idaho Batholith
21
33
35
55
Wyoming
38
39
63
65
Colorado Rocky Mountains
25
42
42
70
1 Calculation based on an assumed F-factor equal to 0.4
Table V.B.6 also provides calculated estimates of the amount of increase in lakewater sulfate
that would be required to acidity the 1st and 5th percentile lake of the subregional ANC
distributions from current values to ANC=0. It was assumed for these calculations that 40% of
the increased SO,2" concentration is neutralized by base cation release (F=0.4) and the remainder
causes a stoichiometric decrease in ANC. If a lower value of F is assumed (for example, F=0.2)
then the estimates of S042" change provided in Table V.B.6 would decrease by 25%. These
calculations suggest that relatively minor increases in lakewater S042" concentration would lead to
chronic acidity (ANC < 0) in the Sierra Nevada and Cascade Mountain ranges. An estimated five
percent of the lakes in these subregions would become acidic with increased SO/" concentration
of only 27 to 30 ueq L'1. This would occur under sulfur deposition loadings of about four times
current levels, based on current SO/' concentrations (Table IV.A.3). Although.uncertainties are
large in current estimates of sulfur deposition in these regions, total sulfur deposition is likely in

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the range of 0.5 to 2 kg S ha"' yr1 (Sisterson et al: 1990). Thus, a reasonable standard for
preventing 5% of the lakes in the Sierra Nevada and Cascade Mountains from becoming
chronically acidic due to sulfur deposition is approximately 2 to 8 kg S ha'1 yr'. In other
subregions of the West, the required S042 increase estimated to cause 5% of the lakes to
become acidic is somewhat higher (55 to 70 ueq L"1), but still low compared to S042"
concentrations currently found throughout the eastern United States. Total sulfur deposition
levels approximately in the range of three times (Colorado) to five times (Idaho) current deposition
would be required to chronically acidify 5% of the lakes in these other western regions. These
estimates equate to acid deposition standards equal to approximately 5 to
10 kg S ha'1 yr '. If we base this analysis on the lowest percentile lake in the subregional ANC
distribution, increased S042" concentrations of 35 to 63 jueq L'1 would cause chronic acidity in the
Idaho Batholith, Wyoming, and Colorado subregions, assuming F=0.4.
MAGIC model projections of change in surface drainage water ANC in response to changes
in sulfur deposition have been shown to be remarkably consistent from region to region in the
eastern United States. Turner et al. (1992) and Sullivan et al. (1992) presented the results of
NAPAP modeling scenarios for 50-year MAGIC simulations for lakes in the Adirondacks, New
England, Mid Atlantic Highlands, and Southern Blue Ridge Province and streams in the Mid-
Atlantic Highlands. Simulations included changes in sulfur deposition (kg ha1 yr ') over 1985
values of -50%, -30%, -20%, 0, +20%, and +30%. Each kg ha1 yr1 change in sulfur deposition
caused approximately a 3.5 weq L'1 change in median lakewater ANC for all regions studied
(Figure IV.E.1). Although the modeled response of individual watersheds to simulated changes in
sulfur deposition was more variable, these results demonstrate that the MAGIC model is strongly
driven by sulfur deposition input values. If we apply these estimates of ANC change (3.5 ,ueq L'1
ANC per kg ha'1 yr1 S deposition) to the measured chemistry of sensitive lakes in the West (Table
IV.A.1); we can obtain an estimate of the amount of sulfur deposition required to drive 5% of the
lakes in the various western regions included in our analyses to chronically acidic conditions

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(ANC < 0). The resulting estimates of sulfur deposition increase are 4 to 5 kg S ha'1 yr"1 for the
Sierra Nevada and Cascade Mountains, and 9 to 12 kg S ha1 yr1 for the other subregions.
Although there are unquantifiable uncertainties associated with such approximations, the results
are generally consistent with calculations based on lakewater sulfate and sulfur deposition values
discussed above. These uncertainties could be substantially reduced by conducting MAGIC
simulations in a suite of watersheds in the western subregions we have identified as potentially
highly sensitive to acidic deposition inputs. Such modeling work has not been conducted.
The estimates of increased S042' concentration required to acidify western lakes within the
lower percentiles of acid-sensitivity, presented above, are based on fall chemistry and chronic
acidification processes. It is likely, however, that sensitive watersheds in the western United
States would experience episodic acidification (especially during snowmelt) at sulfur deposition
levels lower than those that would cause chronic acidification. In most cases, episodic pH and
ANC depressions during snowmelt are driven by natural processes (mainly base cation dilution)
and nitrate enrichments (cf. Wigington et al. 1990). Where pulses of increased S042' are found
during hydrological episodes, they are usually attributable to sulfur storage and release in
streamside wetlands. More often, lake and streamwater concentrations of S042' generally
decrease or remain stable during snowmelt. This is probably attributable to the observation,
based on ratios of naturally-occurring isotopes, that most stream flow during episodes is derived
from pre-event water. Water stored in watershed soils is forced into streams and lakes by
infiltration of meltwater via the "piston effect." This is not necessarily the case for high-elevation
watersheds in the West, however. Such watersheds often have large snowpack accumulations
and little soil cover. Selective elutriation of ions in snowpack can therefore result in relatively
large pulses of both N03 and S042' in drainage water early in the snowmelt. Data supporting the
importance of S042' to spring episodes in the West were presented by Reuss et al. (In Press).
Thus, it appears likely that sulfur deposition will contribute to episodic acidification of sensitive
western surface waters at deposition levels below those that would cause chronic acidification.

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Episodes have been so little studied within the region, however, that it is hot possible to provide
quantitative estimates of episodic sulfur standards for the western subregions of concern.
2) Nitrogen
Streamwater concentrations of N03 are typically below about 5 /xeq L'1 in boreal forested
regions, and such a concentration is considered to have no harmful effect on the biota of
freshwater and near-coastal aquatic systems. Therefore, 5 eq L1 has been suggested as a
reasonable critical concentration for surface waters to protect against significant harmful effects
(Rosen et al. 1992). The relationship between measured wet deposition of nitrogen and
streamwater output of nitrate was evaluated by Driscoll et al. (1989c) for sites in North America
(mostly eastern areas), and augmented by Stoddard (1994). The resulting data showed a pattern
of N leaching at wet-inputs greater than approximately 400 eq ha1 (5.6 kg N ha"'; Figure V.B.1).
Stoddard (1994) presented a geographical analysis of patterns of watershed loss of N
throughout the northeastern United States. He identified approximately 100 surface water sites in
the region with sufficiently intensive data to determine their nitrogen status. Sites were coded
according to their presumed stage of nitrogen retention, and sites ranged from Stage 0 (Figure
III.A.1) through Stage 2 (Figure III.A.3). The geographic pattern in watershed N retention depicted
by Stoddard (1994) followed the geographic pattern of nitrogen deposition. Sites in the
Adirondack and Catskill Mountains, where N deposition is about 11 to 13 kg ha'' yr"\ were
typically identified as Stage 1 or Stage 2. Sites in Maine, where N deposition is about half as
high, were nearly all Stage 0. Sites in New Hampshire and Vermont; which receive intermediate
levels of N deposition, were identified as primarily Stage 0, with some Stage 1 sites (Figure V.B.2).
Based on this analysis, a reasonable threshold of N deposition for transforming a northeastern
site from the "natural" Stage 0 condition to Stage 1 would correspond to the deposition levels
found throughout New Hampshire and Vermont, approximately 8 kg ha' yr'\ This agrees with
Driscoll et al.'s (1989c) interpretation, which suggested N leaching at wet inputs above about

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O
-C
cr
v
W

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> STAGE 0 NO," PATTERN
o STAGE 1 NO," PATTERN
• STAGE 1 N03" PATTERN
AM) INCREASING TREND IN NO"
fc STAGE 2 NO," PATTERN
* STAGE 2 NOg" PATTERN
AND INCREASING TREND IN NO,"

o
2J
Figure V.B.2. Geographic pattern reflected by the stage of nitrogen retention observed at lake
and stream study sites throughout the northeastern United States that have
sufficient data for determining the importance of N03' to seasonal water chemistry
(Source: Stoddard 1994).

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estimates for theAdirondacks (617 eq ha1, Pollack et al. 1989) and both the Bear Brook site in
Maine (310 eq ha'1, Kahl et al. 1993) and Hubbard Brook in New Hampshire (302 eq ha"1,
Stoddard and Kellog 1993).
Lakewater concentrations ot N03' were surprisingly high in many high-elevation sites included
in the Western Lake Survey, despite the possible bias caused by the failure to collect samples at
many of the highest elevation areas due to frozen lake conditions. Based on existing data, it
appears likely that many high-elevation lakes in the West are currently experiencing N deposition
sufficiently high to cause chronic N03 leaching, and likely associated chronic acidification.
Furthermore, it is also likely that many of these sites that exhibit fall concentrations of NOa' in the
range of 10 to 30 ^eq L'1 have substantially higher concentrations during spring. Thus, the weight
of evidence suggest that episodic acidification associated with nitrogen deposition may already
be occurring to a significant degree in many high-elevation western lakes. Unfortunately,
sufficient data are not available with which to adequately evaluate this potentially important issue.
C. UNCERTAINTIES
The major acid deposition issues, problems, and concerns vary from region to region. Sulfur
is of interest for possible chronic acidification in all regions and for episodic acidification primarily
in the West and southeastern Canada. Nitrogen is of interest primarily on an episodic basis in
southeastern Canada, but appears to be of little interest for seepage lakes in the Upper Midwest
and Florida. In the West, nitrogen appears to be an important concern to both chronic and
episodic chemistry. In fact, high-elevation watersheds in the West may currently be experiencing
detrimental levels of N deposition, as reflected by slightly elevated fall N03 concentrations and
high sensitivity.
Just as the key issues and sensitivities vary from region to region, the principal uncertainties
and weaknesses in existing data also vary across the regions selected for study. The most
significant data deficiencies and unanswered questions we have identified are as follows:

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West -	seasonal water chemistry data, especially during spring
deposition data at high elevation sites
biogeochemical data on nitrogen cycling
more detailed analyses of dose/response in individual mountain ranges
Eastern Canada - regional paleolimnological data outside the Sudbury area
regional modeling (i.e., MAGIC)
statistical sampling of chronic lake and stream chemistry
seasonal water chemistry, especially during spring
Upper Midwest - regional paleolimnological data in eastern portion of the region
hydrological and geochemical studies of seepage lakes
Florida -	variable influence of acidic deposition versus groundwater withdrawal in
regulating acid-base status
hydrological and geochemical studies of seepage lakes
In addition to the primary uncertainties outlined above, and discussed in the body of this
report, there are a number of other issues that complicate determination of dose-response
relationships for sulfur and nitrogen. Acidification or recovery of surface water acid-base status in
response to changes in atmospheric deposition can be exacerbated or offset by changes in other
atmospheric factors or land use. Climatic changes (both long-term and inter-annual variability),
forest growth and management, and nitrogen saturation may have relatively large effects that can,
in some cases, overshadow catchment responses to changes in sulfur or nitrogen deposition.
Our ability to incorporate such changes into predictive models is quite limited at present, and
almost no research is being conducted in the United States with which to improve the situation.
The role of trees, forest growth, and historic management can be very important for
determination of acid deposition standards for both sulfur and nitrogen. Trees play a crucial role
in acidification processes by scavenging pollutants from the atmosphere, thereby enhancing
deposition; altering hydrologic flow regimes and evapotranspiration; and taking up base cations
from the soil for tree growth, thereby enhancing soil acidification and base cation depletion
(Jenkins et al. 1990, in press). Any assessment of acid deposition standards, therefore, must be
interpreted within the context of past and future forest management.

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1. Land Use
The influence of historical forest management on the ability of a given forest ecosystem to
process nitrogen is largely unknown. Nevertheless, forest management practices, especially
those that have occurred over many generations, can have important effects on soils (i.e.,
erosion), nutrient supplies (i.e. harvesting), organic material (i.e., litter raking), and thereby many
b aspects of nitrogen cycling and nitrogen effects. For example, by introducing Norway spruce in
high-elevation areas on nutrient-poor soils, forest management in the Vosges Mountains of France
may have exacerbated the impacts of acidic deposition on forests (Landmann 1991). The
introduced Norway spruce has likely contributed to increased dry deposition to the forest and
also increased cation uptake relative to the original forest stands of mixed birch and silver fir. The
observed needle yellowing in Norway spruce in the Vosges Mountains has been attributed to
Mg2+-deficiency. European forests have typically been harvested for many generations, have
been changed in species composition or community type (e.g. conversion from heathland to
forest), and managed or manipulated in a variety of ways. The interactions between these
activities and atmospheric deposition are unknown.
Individual site characteristics and land use history can profoundly influence the response of
watersheds to atmospheric inputs. Such factors seriously complicate the assignment of
standards to prevent achieving critical nitrogen leaching from terrestrial systems. The importance
of land use history is regulating water quality and nutrient cycling has been discussed by Feger et
ai. (1990) for forested sites in Central Europe, which have undergone intense management for
centuries.

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2. Pollutant Interactions
The leaching of N03' from the forest ecosystem in response to nitrogen saturation from
atmospheric inputs implies the concurrent loss of nutrient cations, especially Mg2+ and K+. This
chronic loss of nutrient cations often occurs on sites that
1.	also receive high levels of atmospheric sulfur deposition which is causing cation
leaching, and
2.	are subjected to intensive forest management which also depletes the ecosystem of
nutrient cations in the form of harvested biomass.
Thus, in addition to the interactions between nitrogen deposition and other anthropogenic
activities, other forms of atmospheric pollution are extremely important. The degree to which
concurrent sulfur deposition influences the ecosystem effects of elevated nitrogen deposition is
potentially a major complicating factor in the on-going European research efforts. Although it is
important to attempt to elucidate the nature and magnitude of any such synergistic interactions, it
is also important to evaluate nitrogen effects in the absence of high sulfur inputs. There are
virtually no "pristine" (i.e., low sulfur, low nitrogen deposition) forested areas in Europe. Both
nitrogen and sulfur deposition are substantially above background values virtually everywhere.
This limitation has important consequences for the multi-site international research programs that
are on-going in Europe. There are no suitable pristine control sites available. Sogndal, Norway,
the reference site for NITREX, is relatively pristine, but the vegetation is alpine, rather than forest.
Ballyhooly, Ireland, the reference site for the EXMAN project, receives nitrogen deposition that is
higher than many of the high nitrogen deposition areas in the United States.
Pollutant interactions may be particularly problematic for evaluation of appropriate standards
to protect sensitive western watersheds from episodic effects. Relatively modest increases in both
sulfur and nitrogen deposition can potentially cause significant episodic acidification of low-ANC
lakes, beyond what would be expected from either sulfur or nitrogen alone.
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3. Temporal Considerations
The critical load, as formulated as a science/policy concept in Europe, for atmospheric
deposition of sulfur or nitrogen represents an inherent characteristic of the watershed.
Specification of the critical load, or any kind of acid deposition standard, assumes that the
ecosystem has reached a steady-state with respect to deposition inputs; time scales of
acidification and recovery responses are not considered. Thus, the environmental consequences
of different emission reductions cannot be fully evaluated using only the empirical and steady-
state methods for specifying critical loads (Warfvinge et al. 1992). For example, the long-term
critical load for sulfur at the Birkenes site in southern Norway, required to maintain ANC > 0 (e.g.,
ANC criterion = 0) is estimated to be approximately 50 meq S042" m2 yr"1 (8 kg S ha1 yr"1).
However, the time-dependence derived from the MAGIC model illustrates that to obtain ANC > 0
within 10 years, the target load would be only 1/4 the critical load (12 meq S042' m'2 yr'1); if one
could wait 50 years to achieve ANC > 0, then the target load would be much greater (41 meq
S042" m"2 yr"1) and would approach the long-term critical load (Figure V.B.3, Warfvinge et al. 1992).
Similarly, the starting point can have a very large influence on the model estimate of target load.
Starting with pre-acidification conditions, the MAGIC model estimates that the Birkenes watershed
could tolerate 270 meq S042" m2 yr'1 for ten years before the streamwater would acidify to
ANC = 0. Starting from acidified conditions in 1985, however, MAGIC estimated that the load
would have to be reduced by a factor of 22 (to 12 meq S042"'m"2 yr"1) in order for streamwater to
recover to ANC = 0 (Figure V.B.3, Warfvinge et al. 1992).
Thus, model-based analyses suggest that standards for the protection, or restoration, of
surface water quality must be specified within a temporal context. Standards suitable for
protection of aquatic ecosystems for a short period of time may be less than adequate for long-
term protection. Conversely, reductions in deposition that are insufficient for acidified ecosystem
restoration in the short term may require additional time, rather than additional emissions
reductions, to achieve the desired outcome.

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300
1846 pristine
198S acidified
250
(0 200
O*
V
£ ISO
¦o
(0
o
-J
4)
O)
w
100
£
10 years 50 years 100 years infinito »
critical load
Time Horizon
Figure V.B.3. Target loads of sulfur deposition calculated for the Birkenes catchment in southern
Norway, using a criterion of ANC = 0. Calculations were performed with the
MAGIC model for 10, 50, and 100 years duration, starting from two different initial
states" pristine conditions in 1845 and acidified conditions in 1985. (Source:
Warfvinge et al. 1992)

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VI. SUMMARY AND CONCLUSIONS
There has been a growing international recognition that air pollution effects, particularly from
sulfur and nitrogen, may in some cases necessitate emission controls to reduce atmospheric
deposition. The purpose of this report is to provide technical information required for assessing
the feasibility of adopting one or more acid deposition standards for the protection of aquatic
resources. We have identified four regions for inclusion in this assessment: western and upper
•; midwestern United States, Florida, and southeastern Canada.
Sulfur deposition is a potential concern in ail of the regions under investigation for this report.
Some degree of chronic acidification attributable to sulfur deposition has occurred in southeastern
Canada, in the eastern portion of the Upper Midwest region, and possibly in the Trail Ridge
region of north-central Florida. Regional quantification of the amount of acidification that has
occurred in southeastern Canada and the Upper Midwest is not possible with existing data.
The areas containing low-ANC lakes in the West are confined primarily to the higher elevation
mountainous regions, most of which have been glaciated. There are many low-ANC systems, but
virtually no chronically acidic waters. It is our best professional judgement that a reasonable
standard for protecting sensitive lakes throughout large areas of the West from adverse effects of
chronic sulfur deposition is near 10 kg S ha"' yrIn some mountainous areas of the West,
however, where highly dilute lakes are numerous, such a standard would be considerably lower,
likely in the range of. 2 to 5 kg S-ha'Vyr'1.
Based on analysis of available sulfur dose-response data for sensitive watersheds worldwide,
it is clear that proportional changes in ANC and base cations in drainage waters in response to
changes in sulfur inputs are highly variable. Documented F-factors are generally above 0.5,
although lower values have been found. Perhaps the best available estimate of an appropriate F-
factor for highly sensitive watersheds, such as are found in the western United States, would be
based on the experimental values obtained at Sogndal, in western Norway (F = 0.4). Assuming

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F = 0.4, we calculated that relatively minor increases in lakewater SO/ concentration would lead
to chronic acidity (ANC < 0) in many lakes in the Sierra Nevada and Cascade Mountain ranges.
An estimated five percent of the lakes in these subregions would become acidic with increased
S042' concentration of only 27 to 30 ^eq LAn approximate four-fold increase in sulfur
deposition in these regions, to levels in the range of 2 to 8 kg S ha'1 yr \ would lead to such
increases in lakewater S042' concentrations.
These estimates of increased S042' concentration required to acidify western lakes are based
on fall chemistry and chronic acidification processes. It is likely, however, that sensitive
watersheds in the western United States would experience episodic acidification (especially
during snowmelt) at sulfur deposition levels lower than those that would cause chronic
acidification. Episodes have been so little studied within the region, however, that it is not
possible to provide quantitative estimates of episodic sulfur standards for the western subregions
of concern.
Lakewater concentrations of N03' were surprisingly high in many high-elevation sites included
in the Western Lake Survey, despite the possible bias caused by the failure of EPA's Western
Lake Survey to collect samples at many of the highest elevation areas in the Rocky Mountains
due to frozen lake conditions. Based on existing data, it appears likely that many high-elevation
lakes in the West are currently experiencing N deposition sufficiently high to cause chronic N03'
leaching, and likely associated chronic acidification. Furthermore, it is also likely that many of
these sites that exhibit fall concentrations of NOa' in the range of 10 to 30 ^eq L"' have
substantially higher N03" concentrations during spring. Thus, the weight of evidence suggests
that episodic acidification associated with nitrogen deposition may be occurring to a significant
degree in many high-elevation western lakes. Unfortunately, sufficient data are not available with
which to adequately evaluate this potentially important issue.
The Upper Midwest is characterized by numerous lakes created by repeated glaciations.
Sensitive aquatic resources in the Upper Midwest are largely seepage lakes. Those seepage

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lakes with low base cation concentrations receive nearly all of their hydrologic inputs as
precipitation directly on the lake surface, and have long hydraulic residence times, thus providing
an opportunity for in-lake reduction and assimilation processes to neutralize much of the acidic
inputs. Because concentrations of inorganic nitrogen are uniformly low and snowmelt does not
provide any significant nitrate influx to lakes, the key issue in this region appears to be chronic,
sulfur-driven acidification. The portion of the region most likely to have experienced acidification
from acidic deposition is the Upper Peninsula of Michigan, where acidic seepage lakes are
particularly numerous; acidic deposition is highest for the region, and the lakewater [S042']/[CB]
ratio is commonly > 1.0. In our judgement, a reasonable sulfur standard for the most sensitive
aquatic resources in the Upper Midwest can be approximated by current levels of deposition in
the eastern portion of the region, about 5 kg S ha' yr'V In the western parts of this region,
surface waters are less sensitive to sulfur deposition effects, and an appropriate standard would
be much higher.
Northern Florida contains one of the largest populations of acidic lakes in the United States.
Evidence for acidification of some Florida lakes has been supported by paleolimnological
reconstructions of lake pH, although the case for acidification by acid deposition is equivocal and
the interpretation is complicated by profound regional and local changes in land use and
hydrology. Large groundwater withdrawals of the Floridan aquifer for residential and agricultural
purposes may have contributed to reduced groundwater inflow of base cations into seepage
lakes, thereby causing lakewater acidification. Thus, it is not clear whether current levels of sulfur
deposition have caused recent acidification of lakes in Florida. If such acidification has occurred,
it has likely been restricted to a relatively small geographic area, in the Trail Ridge region.
Beyond the general vicinity of Sudbury, quantitative data on acidification response are scarce
for eastern Canada. Although current chemistry has been investigated to a considerable degree
through lake surveys, the resulting data are insufficient for quantitative dose-response
assessment. We recommend using model-based estimates of appropriate sulfur standards being

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generated by ERL-C (M.R. Church, personal communication) tor the Adirondack region, as a
reasonable first approximation tor southeastern Canada.
Quantification of sulfur and nitrogen dose-response relationships is difficult for the regions
considered in this report (Western and Upper Midwestern United States, Florida, and Eastern
Canada). Limited data availability precludes rigorous quantitative assessment in most cases. In
particular, data are scarce in the following categories:
•	episodic acidification, especially in the West
•	groundwater inflow (and associated neutralization) to seepage lakes
•	seasonal surface water chemistry data, particularly for nitrogen and aluminum
•	model input parameters (especially soils characteristics) for drainage systems
•	deposition (wet and dry) data at high elevation sites
•	regional paleolimnological data, especially in upper Michigan and portions of eastern
Canada

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Appendix A
Assessment Methods tor Eva.uallon o1 Effects on Aquatic Resources Due to
Changes n Atmospheric ~apostion
t. Empirical Data
Curing the past decade, extensive research has been conducted in North America and
Europe concerning the processes that influence watershed sensitivity, surface water chemistry,
and the response of aquatic biota to acidic deposition. Intensive monitoring, field and laboratory
experiments, surveys, trend analyses, and the development and testing of watershed models ail
have provided insight into how aquatic resources respond to changes in the magnitude and
timing of deposition. Each ot these sources o( information has its own strengths and limitations:
the types of data they generate vary, as do the associated uncertainties and assumptions. As a
result, the most defensible conclusions about watershed responses to acidic deposition
incorporate the findings from all of ihese sources, methods, and analytical tools (Sullivan et al.
1992). Consistency of results among methods increases confiderca in the analyses of change;
inconsistencies provide an overaff indication of the degree of uncertainty.
Sensitivity of surface waters to changes in the magnitude and timing of acidic deposition can
be assessed to a limited degree by examination of patterns in current water chemistry.
Differences in surface water chemistry aiong a gradient of low to high deposition may represent
temporal changes in lakes or streams during periods when atmospheric deposition of acids
increased from low to high. In fact, the observed spatial correlation between acidic and low pH
waters and areas of high acidic deposition m the 1970s and early 1980s led, in iarge part, to
formulation of the fundamental hypotheses of surface water acidification and deposition-
walershed interactions. Differences in surface wafer chemistry across gradients in acidic
deposil on do not conclusively cemonstrale a cause-effect relatsonsnip, hcw&yer, because o:her
tactars besides acidic deposition might also vary across the same gradient These other factors

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could include, for example, inherent watershed sensitivity, marine influences, human disturbances,
or land use. Nevertheless, the distribution of current chemistry across gradients in deposition
provides useful information, and proportional changes across space among the various ions
provide an estimate (albeit inconclusive) of the magnitude of past acidification.
Measured changes in water chemistry potentially can provide the most direct and most
certain evidence of acidification. Reliable long-term (> 50 years) historical data are seldom
available, however. Because of large uncertainties and poor documentation of the sampling and
analytical procedures used in the past, comparisons between historic and recent chemical
measurements must be interpreted with caution. More recent chemical monitoring data, generally
for up to one or two decades, are not subject to the same methodological limitations as the
"historical" data. Often the trends, if they exist, are too small to be separated from natural
variability in short-term data sets, however. Nevertheless, sulfur deposition has changed
markedly over the period oj measurement in some areas, and quantitative information can be
obtained. These data are of great value because the changes have actually been measured, and
data interpretation is not constrained by the various assumptions inherent in the use of other
assessment techniques.
2. Paleolimnoloqical Data
Chemical inferences based on fossil remains preserved in lake sediments of freshwater
algae, particularly diatoms and chrysophytes, are commonly used to quantify past acidification.
The fossil remains of these organisms are good indicators of past takewater chemistry because
many species occur in large numbers and they often have narrow ecological (water chemistry)
tolerances. Paleolimnology involves quantitative reconstruction of past lakewater chemistry,
based on relationships between relative abundances of algal taxa and current chemistry. The
predictive relationships are then applied to algal data from one or more cores of sediment
collected from each study lake. Individual sediment layers are analyzed and the date when that

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increment was deposited in the lake bottom is estimated using radioisotopes. Paleolimnological
reconstructions of lakewater acid-base chemistry have been developed for many hundreds of
lakes in the United States, Canada, and Europe during the last decade (Sullivan 1990). Such
data provide an independent means of evaluating (verifying) the accuracy of model projections of
acid-base chemistry (e.g., Wright et al. 1986, Jenkins et al. 1990, Sullivan et al. 1992).
3.	Experimental Data
Experimental manipulations of the acid-base chemistry of whole lakes or whole catchments
provide very valuable measurements of acidification and/or de-acidification (recovery).
Interpretation of results of a lake acidification experiment is somewhat limited because the
manipulation pertains only to the lake itself. Results reflect in-lake processes, but do not account
for processes that occur within the watershed. Results of lake manipulations do provide useful
quantitative information regarding in-lake neutralization of.acids, however. Whole catchment
manipulation studies have been conducted throughout Europe since 1984 (see, e.g., Sullivan
1993). Many of these experiments provide quantitative data which are useful in comparison with
model projections, and can be evaluated relative to similar systems in North America.
4.	Models
The principal models used within NAPAP for projecting acidification response were the
MAGIC model for drainage systems and the IAG model for seepage lakes. MAGIC and IAG, like
other processed-based models, are simplified representations of catchment and in-lake
processes. Although rooted in hydrochemical principles, the models include major temporal and
spatial process aggregation, and seme catchment processes are not well represented. Relatively
little attention was placed on model evaluation and validation within the NAPAP program, largely
because of time constraints relative to the delivery date of the 1990 Integrated Assessment. In

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the absence of such validation, however, future scenario projections are subject to large, and
unquantifiable, uncertainties.
The most extensive model validation exercise conducted for a process-based acid-base
chemical model was the recent comparison between MAGIC model hindcasts and
paleolimnological inferences of historical acidification for a set of 33 statistically-representative
Adirondack lakes (Sullivan et al. 1991, 1992, in review). Both assessment methods demonstrated
acidification of low-ANC Adirondack lakes since pre-industrial times. They differed primarily in
that MAGIC inferred greater acidification and also that acidification had occurred in all lakes in the
comparison. The diatom approach inferred that acidification had been restricted to low-ANC
lakes (< about 50 ,ueq L"1).
The lack of organic acid.representation in the MAGIC simulations conducted by the U.S.
EPA's Direct Delayed Response Project (Church et al. 1989), Sullivan et al. (1991), and the
Integrated Assessment analyses was identified as an important factor contributing to the observed
discrepancy. Organic acids often exert a large influence on surface water acid-base chemistry,
particularly in dilute waters. Sullivan et al. (in review) investigated the potential role of organic
acids in influencing the comparison results, and incorporated an organic acid representation
developed by Driscoll et al. (1994) into the MAGIC model. The revised model provided hindcast
estimates of pre-industrial lakewater chemistry that more closely matched diatom-inferred
estimates for the same, lakes (Sullivan et al., in review). The median F-factor, where:
F = A CB -f A (SO,2" + N03 )	(1)
calculated by MAGIC for historic acidification was 0.56. The 5th and 95th percentiles of the F-
factor distribution were 0.25 and 0.67, respectively. MAGIC projections of future acid-base
chemistry in the year 2034, under a scenario of 50% reduction in sulfur deposition, yielded
estimated F-factors somewhat higher, with a median of 0.73 and 5th and 95th percentiles of 0.39
and 0.80, respectively (Sullivan, T.J., unpublished).

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a. Static Models and Empirical Approaches
The empirical method of choice at present in Europe for quantifying critical loads of sulfur for
aquatic systems is the Henriksen et al. (1990a,b; 1992) steady state approach which is based on
two equations. The first equation estimates the pre-industrial (or initial) base cation concentration
[6C]0, in lakewater as:
[BC]0 = [BC], - F * ([SO„]t - [S04]0)	(2)
where all constituents (in fieq L'1) are sea salt corrected and the subscripts O and t pertain to
initial and current conditions, respectively. The F-factor is defined as the change in base cation
concentration divided by the change in S04 concentration, or [S042" + N03] where NO3" is
important, (Equation 1) and must be derived using some other means. Equation 3 is the actual
critical load (CL) calculation:
CL = Q * ([BC]0 - (ANC]limil) - BCd	(3)
where Q is runoff. ANClim„ is the critical ANC level below which the waters are desired not to be
acidified, and BCd is the nonmarine base cation deposition.
There are important difficulties associated with the failure of Equation 3 to include
incorporation of the F-factor or enhanced weathering from acidic deposition, the failure to
consider nitrogen, and the uncertainties associated with estimation of the F-factor used in
Equation 2. It has also become apparent that defining the critical load for surface waters as the
leaching of alkalinity or flux of base cations, using a steady-state approach, tends to overestimate
the sensitivity of humic waters (e.g. Forsius et al. 1992). The ANC,,mit should be a function of
natural organic acidity, as well as mineral acidity. For example, Forsius et al. (1992) calculated
critical loads of sulfur using the steady-state approach for a subregion of Finland as

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approximately 50 meq/m2/yr (8 Kg S ha1 yr'). They noted, however, that the current deposition
exceeds this amount in large areas, yet evidence of regional impacts on fisheries has not been
found.
The approach described by Henriksen calculates critical loads on the basis of base cation
fluxes into and from catchments. It involves the following steps:
1.	Adjust surface water concentrations of sulfate and base cations for marine deposition,
generally assuming that all CI' in surface water originates from sea salt deposition.
2.	Estimate the pre-acidification nonmarine base cation concentrations ([BC]o) from the
current measured concentration ([BC],"), assuming values for the F-factor
(A [BC]'/A [S04]') and pre-industrial nonmarine sulfate concentration [S04£. Henriksen
estimates [S04£ as a function of [BC],'. Alternatively, [S04]g can be assumed equal to a
low value, for example 0 to 15 /zeq L"'.. Henriksen proposed estimating the F-factor as
F = sin (90 x [QC]'JS)	(4)
where the variable S is the value for [BC]* above which F=1. A value of 400 ^eq L"' was
specified for S for Norwegian lakes.
3. Calculate the critical load (CL) as the difference between the base cation flux out of the
system (expressed as "excess" above the critical ANC limit) and the base cation flux into
the system from deposition, using Equation 3.
Bernert and Sullivan (1990) examined the effect of varying the model input specification for
the [SOJo estimate. An assumed value of 20 ueq L'1 was compared with the method proposed
by Henriksen, which assumed a relationship between [BC]* and [SOX The results were very
similar, suggesting that differences in the background sulfate assumption do not have a
substantial effect on results of the calculation. This is reasonable because A [SO„]' only enters
into the calculation for estimation of [BC]^, and it is modified by the F-factor fraction. The critical
load calculation (Equation 3) implicitly assumes that the pre-industrial nonmarine mineral acid
anion concentration was zero and that F=0, because the base cation flux out is based on the
difference between pre-industrial base cation concentration and the ANC limit. It also does not
consider the possibility of changes in base cation deposition in conjunction with changes in sulfur

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deposition. The magnitude of the effects of these assumptions on the resulting critical load
calculations are not known.
Using the empirical model of Henriksen et al. (1992), Jenkins et al. (in press) estimated that
85% of the lochs in the Galloway region of Scotland would be protected (based on critical ANC of
0) with a maximum sulfur deposition of about 20 meq/m2/yr (3.2 Kg S ha'1 yr"1).
The relationship between Henriksen's estimates for the F-factor and [BC]t* is presented in
"Figure A.1. Also presented are the estimates for the F-factor derived from diatom-inferred
concentrations for Adirondack Mountain lakes (Sullivan et al. 1990a). Paleolimnological estimates
of the F-factor are higher than Henriksen's estimates for Norway. It should be noted, however,
that there is no a priori reason to expect them to be the same. Estimates of acidification
response using the Henriksen approach will imply greater sensitivity to the effects of sulfur
deposition than were observed for Adirondack lakes. It is not unreasonable to assume, however,
that Norwegian catchments may be inherently more sensitive (e.g...Wright 1988).
Selection of the ANClimit has a major influence on critical load calculations for surface waters
in some areas. Henriksen et al. (1990a) produced a map of the estimated critical loads for
Norway using an ANClimit of 0 ,weq L'\ largely because Henriksen and co-workers assumed that
many Norwegian lakes had ANC close to zero in their natural state. An ANC,imit of 50 fieq L'1 has
been suggested for Sweden, because this is the typical target alkalinity of liming programs.
Norway appears to be the European country most sensitive to acidic deposition. Setting an
ANClimil of 50 /ieq L' for Norwegian lakes results in most of southern Norway having critical load
estimates of zero (or less) (Henriksen et al. 1990b). Henriksen et al. (1990b) suggested an
ANClimrt of 20 ^eq L"' for evaluation of critical loads for the protection of freshwater fish in Norway.
Using this criteria, Henriksen et al. (1990b) illustrated the importance of the selection criteria for
designating the percentage of lakes that must exceed the criteria for the purpose of mapping
critical loads. For example, if the most sensitive lake in each map grid in Fennoscandia is used
for mapping the critical load, then the resulting critical load of S is less than 50 meq/m2/yr (8 Kg S

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Diatom Inferred Spline
Henriksen Equation
rc 0.8
100	200	300	400	500
Base Cations (qeq/L)
600
Figure A.1. Estimated F-factor according to Henriksen's steady-state water chemistry method for
southern Norway, expressed as a function of current lakewater base cation
concentration (sea salt corrected). Also presented are diatom-inferred F-factors and
base cation levels for Adirondack lakes, estimated by Sullivan et al. (1990a).
Paleolimnologica! study lakes identified as having watershed disturbances are coded
by open circles. Relatively undisturbed systems are coded by filled circles.

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ha'1 yr') for more than 80% of the area mapped. Using the 25th percentile as representative for
each map grid, only about 55% of the area had an estimated critical load less than 50 meq/m2/yr
(Henriksen et al. 1990b).
b. Dynamic Models
The principal model used for acidic deposition policy analyses in the United Sates (e.g.
NAPAP 1991) has been the Model for Acidification of Groundwater in Catchments (MAGIC)
(Cosby et al. 1985a,b). MAGIC is also commonly used for calculating critical loads to soils and
surface waters in Europe (e.g., Jenkins et al. in press). The major processes included in the
model are deposition, sulfate adsorption, cation exchange, C02 dissolution, aluminum dissolution
and precipitation, weathering, uptake and release of cations by vegetation, and export in runoff.
The input data required by MAGIC include wet and dry deposition of major ions, average
catchment soil parameters (depth, bulk density, cation exchange capacity, exchangeable base
cations), runoff volume, and net uptake of major ions by the vegetation. All input data are
required for the calibration year and estimates for the past history are also needed.
Jenkins et al. (in press) used the steady state empirical model of Henriksen et al. (1992) and
the MAGIC model to estimate the critical load of sulfur for 38 lochs in the acid-sensitive Galloway
area of Scotland. MAGIC also was used to estimate the influence of afforestation on the critical
load calculations. Across the range of study sites, the agreement between the two approaches
was reasonably good, although at the most sensitive end of the distribution, MAGIC estimated
consistently lower critical loads. Critical loads calculated by MAGIC for soils were about three
times higher than for surface waters at a given site (Jenkins et al. in press). This suggests that
sulfur deposition control strategies that protect surface waters will also protect soils. This may not
be the case for nitrogen, however.
Processes that regulate the nitrogen cycle are not well quantified, thus precluding the use of
a rigorous dynamic model to estimate N critical loads at this time. Until it becomes possible to

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parameterize a dynamic N model, a simple mass balance model for N is probably the best
available approach.

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