AQUATIC EFFECTS OF ACIDIC DEPOSITION - CHAPTER 5 pdf

Diverse data are available from a variety of sources with which to quantifythe watershed acidification response, as well as recovery from acidification.Such data shed light on the sensit

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5

Chemical Dose–Response Relationships and Critical Loads

5.1 Quantification of Chemical Dose–Response Relationships

There has been a growing international recognition that air pollution effects,particularly from S and N, may in some cases necessitate emission controls toreduce or limit future increases in atmospheric deposition Measures toreduce emissions must rely on known or estimated dose–response relation-ships that reflect the tolerance of natural ecosystems to various inputs ofatmospheric pollutants This need has stimulated interest in evaluating theefficacy of establishing one or more standards for acid deposition The CleanAir Act Amendments of 1990 (CAAA) also included requirements to assessthe effectiveness of the mandated emissions controls via periodic assess-ments, and to submit an EPA report on the feasibility of adopting one or moreacid deposition standards to Congress

Diverse data are available from a variety of sources with which to quantifythe 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 studiesthat have been conducted are complicated by different relative watershedsensitivities, S deposition loading rates (and changes in those rates), the rela-tive importance of N leaching and N saturation, temporal considerations, andnatural (especially climatic) variability In addition, these quantitative datahave been generated in vastly different ways, including monitoring, space-for-time substitution, whole-watershed or whole-lake acidification, whole-watershed acid exclusion, paleolimnology, and modeling The only way inwhich different approaches can be compared on a quantitative basis is by nor-malizing surface water response as a fraction of the change in SO42- concentra-tion (or SO42- + NO3- concentration where NO3- is also important) Theprincipal ions that change in direct response to changes in (SO42- + NO3) con-centration are ANC (which can be expressed as [HCO3- - H+]), base cations(CB), inorganic aluminum (Ali), and organic acid anions (A-) The proportional1416/frame/C05 Page 115 Wednesday, February 9, 2000 2:09 PM

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116 Aquatic Effects of Acidic Deposition

changes in (HCO3- - H+), Ali, CB, and A- concentrations should sum to 1.0 inorder to satisfy the electroneutrality condition For aquatic systems that arerelatively insensitive to acidic deposition, ∆CB approximates ∆(SO42- + NO3-),and the F factor (Henriksen, 1982) approximately equals 1.0:

~

where brackets indicate concentration in µeq/L and changes in other uents are insignificant Where acidification occurs in response to acidic dep-

percentage of the overall surface water response and, therefore, the F factor

is less than 1.0 The F factor is important in evaluating criteria for establishingacid deposition standards because it provides the quantitative linkagebetween inputs of acid anions (e.g., SO42-, NO3-) and effects on surface waterchemistry An important limitation of the F factor concept, however, is thatthe value of F is likely to change as the base cation pools in watershed soilsbecome depleted by acid deposition inputs

Quantitative dose–response relationships for S have been determined,using a variety of approaches, in a number of regions in North America andEurope Such studies have included, for example, measured changes in waterchemistry during periods when S deposition changed appreciably, regionalpaleolimnological (e.g., diatom-inferred change in pH and ANC) investiga-tions, whole-catchment manipulation studies, and intensive process model-ing Each type of study has provided quantitative estimates of dose–responsethat entail different sets of assumptions and limitations Taken together, theyprovide a good indication of the range of quantitative acidification response

As a result of these recent studies, we are much better able to quantify fication and recovery relationships than we were in 1990

acidi-5.1.1 Measured Changes in Acid–Base Chemistry

Measured changes in surface water chemistry in areas that have experiencedshort-term (less than 20 years) changes in chemical constituents in response tochanges in mineral acid inputs are available from a number of sources Propor-tional changes in ANC, base cations, and Ali relative to changes in SO42- or(SO42- + NO3-) concentrations were summarized by Sullivan and Eilers (1994)for lakes and streams in which such changes had been measured Theyincluded lakes in the Sudbury region of Ontario, the Galloway lakes area ofScotland, a stream site at Hubbard Brook, NH, and catchment manipulationexperiments in the RAIN project in Norway and Little Rock Lake in Wisconsin.Most of the observed changes were coincident with decreased acidic deposi-tion, and it is unclear to what extent acidification and recovery are symmetri-cal F-factors in the range of 0.5 to 0.9 are apparently typical for lakes havinglow base cation concentrations, although lower values (0.35 to 0.39) were

SO42-+NO3

∆ -

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Chemical Dose–Response Relationships and Critical Loads 117

TABLE 5.1

Measured Short-Term Changes in Surface Water Chemistry Associated with Changes

e Little Rock Lake experiment involved manipulation of lake only.

∆ HCO 3 -

- ∆Al

∆SO 4 2-

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observed for the highly sensitive catchments at Sogndal, Norway that are acterized by thin soils and much exposed bedrock, as is common in many areas

char-of southern Norway and the western U.S The proportional change in ANC ative to the change in (SO42- + NO3-) was variable, within the range of 0.1 to 0.5(Table 5.1) The proportional change in Al was smaller, ranging up to 0.15.These measured values of acidification and deacidification change in ANC and

rel-Al are somewhat smaller than previously anticipated

Relatively early in the international efforts to quantify the acidification

would be in the range 0 to 0.4 More recent research (e.g., Table 5.1) hasshown this earlier estimate to be too low in most cases Based on measured

factors below 0.4

TABLE 5.2

Inferred Long-Term Regional Changes in Surface Water Chemistry Associated with Estimated Changes in Mineral Acid Anion Concentrations, Using the Technique of Space-for-Time Substitution

NE U.S 0.13 0.54 0.07 Sullivan et al.,

1990a

Analysis restricted to lakes having current ANC ≤ 25 µ eq/L

TABLE 5.3

Diatom-Inferred Long-Term Changes in Lake-water ANC as a Fraction of

Region

Number

Adirondacks, NY 48 0.11 Sullivan et al.,

1990a

Statistical sampling Adirondacks, NY 25 0.18 Sullivan et al.,

1990a

Acidic lakes only a

Northern New England 12 0.30 Davis et al., 1994 Lakes were selected

that were presumed

to be acid-sensitive Florida (Lakes Barco,

- ∆C B

∆SO 4 2-

- ∆Al

∆SO 4 2-

-∆ANC

∆SO 4 2-

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In addition to the measured acidification and recovery data presented in

Table 5.1, 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 5.2), diatom-inferences of historical acidification (Table 5.3), and results

of process-based model hindcasts or future forecasts (Table 5.4) Each of these

methods has its own assumptions and limitations, and none are as robust as

results of actual field measurements of response Major advantages of these

alternative sources of quantitative data, however, are that they primarily reflect

acidification, rather than recovery, scenarios, and that they sometimes include

longer periods of response than do the available direct measurements

Results of space-for-time substitution must be interpreted with caution This

approach is based on the assumption 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

dep-osition (e.g., soil characteristics, land use impacts) do not co-vary with

depo-sition Results should therefore be considered only semiquantitative

Nevertheless, available data using this method (Table 5.2) appear similar to

results of measured values shown in Table 5.1

The spatial distributions of lake-water chemical variables across a

longitu-dinal gradient in the upper Midwest for low-ANC groundwater recharge

TABLE 5.4

Projections of Acidification or Recovery Responses

Number

of Lakes

or Streams

50% reduction

in S deposition

33 0.73 0.39 Sullivan,

unpublished Wilderness

lakes, Western

U.S.

Forecasted 3-fold increase

in S deposition

15 0.34 0.03 Eilers et al.,

1991 Bear Brook, ME Response to

experimental watershed acidification

1 0.85 — Norton et al.,

1992

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seepage lakes (Figure 4.5) provides a good example of the use of

space-for-time substitution to evaluate acidification dose–response relationships

These distributions also 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

sys-tems of most interest, however, it is difficult to substantiate in terms of

mag-nitude 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, and hydrologic retention times are long Sulfur

deposition appears to be of greater importance, and potential chronic effects

are of greater interest than episodic effects because of the nature of the

hydrology of sensitive resources in the region Based largely on the results of

space-for-time substitution analyses, Sullivan and Eilers (1994) concluded

that current deposition in the eastern portion of the region (approximately 5

kg S/ha per year) is a reasonable approximation of the deposition level

required to protect the most sensitive aquatic receptors Resources in the

western portion of the region are less sensitive, however, and an appropriate

standard for S deposition would be much higher Because S deposition has

been decreasing in recent years, it does not appear that acidic deposition is an

important environmental concern in the upper Midwest at this time

An S 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

Con-trol Agency (MPCA) to identify natural resources within the state that were

threatened by acid deposition and to develop both an acid deposition

stan-dard and an emissions control plan Small, poorly-buffered lakes in

northcen-tral and northeastern Minnesota were identified as the resources at greatest

risk Based on model simulations, MPCA selected a threshold pH for

precip-itation of 4.7, below which damage to aquatic biota was thought to occur with

prolonged exposure This threshold pH was correlated with SO42- deposition

data, and a standard was determined that allowed no more than 11 kg/ha of

wet SO42- to be deposited during any 52-week period (3.7 kg S/ha per year)

(MPCA, 1985) This standard is fairly stringent In fact, 6 of 12 monitoring

sites in Minnesota exceeded the standard in 1992 (Orr, 1993) There appears

to be a limited scientific basis for such a standard for protection of aquatic

resources in Minnesota

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 acidification

in recent decades (Sweets, 1992) Proportional changes in diatom-inferred

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pre-industrial times show estimates ranging from 0.1 to 0.3 (Table 5.3), in close

agreement with measured values (Table 5.1)

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, 1992), 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

Lyse-vatten 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

sed-imentary 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

dea-cidifying Round Loch of Glenhead, Scotland to be somewhat smaller than the

measured pH recovery since the late 1970s pH reconstructions from the

sed-iment cores showed an average recovery of 0.05 pH units Measured

increases in pH between 1978–1979 and 1988–1989 averaged 0.23 pH units

The authors attributed this difference to attenuation of the reconstructed pH

record owing to sediment mixing processes

Dixit et al (1992) analyzed sedimentary diatoms and chrysophytes from

Baby Lake (Sudbury, Ontario) to assess trends in lake-water chemistry

asso-ciated with the operation, and closure in 1972, of the Coniston Smelter

Extremely high S emissions caused the lake to acidify from pH

approxi-mately equal to 6.5 in 1940 to a low of 4.2 in 1975 Following closure of the

smelter, lake-water 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

habi-tats where pH changes may be buffered by chemical and biological

pro-cesses Alternatively, such an attenuation could be a result of sediment

mixing processes

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Some upper Midwestern lakes have acidified since pre-industrial times

However, based on available paleolimnological data, there is little

paleolim-nological evidence suggesting that widespread acidification has occurred in

this region (Kingston et al., 1990; Cook et al., 1990) Land use changes and

other human disturbances of upper Midwestern lakes and their watersheds

have probably exerted more influence on the acid–base chemistry of lakes

than has acidic deposition (Eilers et al., 1989a; Kingston et al., 1990; Sullivan,

1990) This is because acidic deposition has occurred at a much lower level in

the upper Midwest than in most areas of the eastern U.S 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., 1990b), acidic deposition is highest for the region, and

the [SO42-]/[CB] ratio is commonly greater than 1.0 (Figure 4.5) The

percent-age 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., 1988b),

which is comparable to heavily impacted areas of the Northeast

Diatom-inferred pH data are available for only two lakes in upper

Michi-gan, McNearney and Andrus Lakes McNearney Lake was naturally acidic

prior to this century and is therefore 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

acidifi-cation, although quantitative data are lacking regarding the amount of

acid-ification that occurred in the past or the dose–response relationships of these

systems In addition to the scarcity of paleolimnological data within the

por-tion of the upper Midwest most likely to have experienced widespread

his-torical acidification, there is also a paucity of basic biogeochemical data on

the response of the predominant lake type in this region to atmospheric

inputs of S and N

Historical changes in Florida lake-water chemistry, as inferred from

dia-toms, showed a distinct geographical pattern All five of the

paleolimnologi-cal 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

No clear evidence of acidification was observed for lakes in the 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

sev-eral 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

Pan-handle, 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

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(ca 1 to 3% of total hydrologic budget; cf Baker et al., 1988b) Groundwater

monitoring data collected adjacent to Lake Five-O suggested, 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 Cl

-balances for Panhandle lakes also suggested substantial groundwater

inflows, ranging from 10 (Moore Lake) to 29% (Lofton Ponds) (Pollman and

Sweets, 1990)

Superimposed on the complex heterogeneity of Florida lakes is a high

inci-dence of anthropogenic disturbance Of the 159 total lakes sampled by ELS-I

in Florida, all but 37 were judged by Baker et al (1988b) to have substantial

shoreline or watershed disturbances, mostly related to agriculture Besides

the increased atmospheric deposition in Florida in the 1950s, other changes

have also occurred The human population has increased markedly, as has

freshwater withdrawal from the Floridan aquifer (Aucott, 1988) As a result,

the potentiometric head has declined substantially in the Trail Ridge area

(Healy, 1975; Aucott, 1988; Pollman and Canfield, 1991) The effects of water

withdrawal on the acid–base status of lakes is not well understood

For undeveloped lakes in the northcentral peninsula, lake-water

chemis-try is consistent with an hypothesis of acidification by acidic deposition

(Hendry and Brezonik, 1984; Eilers et al., 1988c; Baker et al., 1986, 1988b)

Evaporative concentration of modest amounts of acidic deposition, and

in-lake retention of SO42- and NO3- appear to be important processes However,

Eilers et al (1988c) concluded it is unlikely that the mechanisms of

acidifi-cation of clearwater lakes in Florida and the linkages to atmospheric

depo-sition 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 lake-water chemistry in these

precipi-tation-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 Trail Ridge district may be

attrib-utable to acidic deposition, as compared to other anthropogenic activities,

especially groundwater withdrawal

5.1.4 Model Estimates of Dose–Response

U.S., using the MAGIC model, show reasonably close agreement with

mea-sured F-factors for acid-sensitive systems (Tables 5.1 and 5.4)

Model-gen-erated 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

reason-ably comparable to the measured values at the highly sensitive Sogndal site

(0.35 to 0.39) MAGIC forecasts for western lakes, however, yielded

esti-mated F-factors that were substantially lower (median 0.34, 5th percentile

0.03; Table 5.4) It is not clear how representative these forecasts might be

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for western lakes, in general, or how accurate the estimates are for the eled lakes Nevertheless, these comparative data suggest that western sys-tems are as sensitive, or perhaps more sensitive, than any of the watershedsfor which acidification and/or recovery responses have been more rigor-ously quantified

mod-5.2 Critical Loads

It has been well documented that acidic deposition has caused tal degradation of surface waters, soils, and forests in certain areas Such deg-radation has been more widespread in Europe than in North America, owingpartly to the fact that many regions of Europe have received much higherdeposition of S and N for a longer period of time than have comparableNorth American ecosystems Recent emissions control efforts have focused

environmen-on attempts to reduce depositienvironmen-on sufficiently to permit ecosystem recovery, ifnot to pre-acidification levels, at least to ecologically acceptable levels Thekey questions facing scientists and policy-makers, therefore, have to do withthe degree in space and time to which S and N emissions will need to bereduced in order to allow ecosystem recovery to proceed (Jenkins et al., 1998) Public policy measures to reduce emissions must be based upon quantifieddose–response relationships that reflect the tolerance of natural ecosystems

to various inputs of atmospheric pollutants This need has given rise to theconcepts of critical levels of pollutants and critical loads of deposition (e.g.Bull, 1991, 1992), as well as interest in establishing one or more standards foracid deposition A critical load can be defined as “a quantitative estimate of

an exposure to one or more pollutants below which significant harmfuleffects on specified sensitive elements of the environment do not occuraccording to present knowledge” (e.g., Nilsson, 1986; Gundersen, 1992) Such

an approach to establishing a standard is intuitively satisfying However, theassignment of a standard or critical load of S or N for any particular regionmay be difficult to defend scientifically A variety of natural processes andanthropogenic activities affect the acid–base chemistry of lakes and streams,

in addition to atmospheric deposition of S and N The loadings of N or S 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 bedifficult to quantify

The basic concept of critical load is relatively simple, as the threshold centration of pollutants at which harmful effects on sensitive receptors begin

con-to occur Implementation of the concept is, however, not at all simple orstraightforward Practical definitions for particular receptors (soils, fresh

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waters, forests) have not been agreed to easily Different research groups haveemployed different definitions and levels of complexity (Bull, 1991, 1992).Constraints on the availability of suitable, high-quality regional data havebeen considerable

The acid–base chemistry of surface waters typically exhibits substantialintra- and interannual variability Seasonal variability in the concentration

of key chemical parameters often varies by more than the amount of fication that might occur in response to acidic deposition Such variabilitymakes quantification of acidification and recovery responses difficult, andalso complicates attempts to evaluate sensitivity to acidification basedsolely on index chemistry, as is typically collected in synoptic lake or streamsurveys Seasonal variability is particularly problematic in the assessments

the year 1993 Interestingly, the U.K was widely criticized for failing to signthe First Sulfur Protochol and thereby joining the “30% Club,” and yet subse-quently agreed in 1994 to an 80% reduction in SO2 emissions by the year 2010.This is indicative of the fact that enormous political changes have occurredsince the 1980s We scientists like to believe that those political changes havebeen the direct result of our scientific advancements

The majority of the critical loads work to date has been conducted inEurope A number of documents have been prepared in conjunction with theUN/ECE critical loads research efforts over the past decade These haveincluded documentation of methodologies (e.g., ECE , 1990) and presentation

of critical loads maps for portions of Europe In addition, a number of otherbackground documents have been prepared in conjunction with the ongoingcritical loads research efforts in Europe (e.g., Gundersen, 1992; Kämäri et al.,1993; Hessen et al., 1992; Lövblad and Erisman, 1992)

A simplistic and generalized attempt to quantify critical loads for S and

N was presented at the Skokloster workshop (Nilsson and Grennfelt, 1988),based on a long-term mass-balance approach A stable base cation pool wasused 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 NO3- and SO42- above the production rate of basecations via weathering will eventually lead to soil acidification The permis-sible input of N for designation of the critical load was the amount allocated

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to forest growth, forest floor accumulation, and an acceptable leaching of 1

to 2 kg N/ha per year On this basis, Nilsson and Grennfelt (1988) estimatedcritical loads of N for Europe to be in the range of 3 to 20 kg N/ha per year,depending on forest productivity

Although some ECE working groups have developed fairly complex, cess-based approaches, the severe constraints on data availability generallynecessitate creating maps based on the more simplistic steady-stateapproaches that tend to have more substantial problems For example, a cal-culation frequently employed for estimation of the critical load of S to surfacewaters is based on assumed pre-industrial and current base cation fluxes(Henriksen et al., 1990a,b, 1992) There is significant uncertainty in the esti-

pro-mates of current base cation input, 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 beenassumed that it will not be possible to reduce loads below critical values forsome sensitive systems in Europe, and also that dynamic watershed pro-cesses cause lag periods in the acidification and recovery responses Theseproblems have given rise to the concept of target loads (e.g., Henriksen andBrakke, 1988) that implies policy relevance, rather than strictly ecological jus-tification Critical loads and target loads are conceptually different A criticalload is a characteristic of a specific environment that can be estimated by avariety of mechanistic and empirical approaches A target load can be based

on political, economic, or temporal considerations, and implies that the ronment will be protected to a specified level (i.e., certain degree of allowabledamage) and/or over a specified period of time For example, a given targetload may be sufficiently low as to protect a particular ecosystem from signif-icant environmental degradation over a 10-year period but, in fact, may besubstantially higher than would be required for long-term protection of thatecosystem There has been a rapid acceptance of the concepts of critical andtarget loads throughout Europe for use in political negotiations concerningair pollution and development of abatement strategies to mitigate environ-mental damage (e.g., Posch et al., 1997)

envi-Criteria of unacceptable change used in critical loads assessments are ically set in relation to known effects on aquatic and terrestrial organisms Forprotection of aquatic organisms, the ANC of runoff water is most commonlyused (Nilsson and Grennfelt, 1988; Henriksen and Brakke, 1988; Sverdrup etal., 1990) Critical limits of ANC, that is, concentrations below which ANCshould not be permitted to fall, have been set at 0, 20, and 50 µeq/L for vari-ous applications (e.g., Kämäri et al., 1992) Designation of an ANC limit isconfounded, however, by natural acidification processes that can also reduceANC to low, or even negative, values

typ-An ANC limit of 0 has been adopted by the U.K for the national mapping

of critical loads for surface waters (Harriman et al., 1995a) This has beendefined as the ANC at which there exists a 50% probability of survival ofsalmonid fisheries (Sverdrup et al., 1990) However, recent evidence suggeststhat, for Scottish fisheries, sites with mean surface water ANC less than or

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