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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'
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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0.73 0.39 Sullivan, unpublished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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