35

3

Chronic Acidification

Chronic acidification of surface waters refers to loss of ANC or reduction in
pH on a chronic, or annual-average, basis. Chronic acidification is often
evaluated by studying changes in surface water chemistry during periods
when that chemistry is expected to be relatively stable. These are generally
summer or fall for lakes and spring baseflow (in the absence of storms) for
streams. Attempts to measure chronic acidification focus to some extent on
a moving target. Lake-water chemistry tends to be relatively stable during
summer and fall, compared to other times of the year, as does spring base-
flow chemistry in streams. There is still, however, often significant variabil-
ity in that chemistry. Water chemistry exhibits changes on both intra and
interannual time scales in response to a host of environmental factors. Key
in this regard are short-term and long-term climatic fluctuations that gov-
ern the amount and timing of precipitation inputs, snowmelt, vegetative
growth, depth to groundwater tables, and evapoconcentration of solutes.
Many years of data, therefore, are required to establish the existence of
trends in surface water chemistry, much less assign causality to changes
that are found to occur.
There have been many advancements in the scientific understanding of
chronic surface water acidification since 1990. Several studies that had been
initiated during the original NAPAP research effort were completed post-
1990 and research results from those programs continue to be published. A
major research effort was conducted in Europe regarding the dynamics of N-
driven acidification and related processes in both terrestrial and aquatic eco-
systems. New predictive models have been developed and some previously

of available water chemistry data (e.g., Omernik and Powers, 1982; Eilers and
Selle, 1991) refined and expanded this image of sensitive areas in North

FIGURE 3.1

Major areas of North America containing low-ANC surface waters as defined by Charles (1991).

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Chronic Acidification

37
America. The extensive research programs conducted in Europe, Canada,
and through NAPAP provided additional insight into factors contributing to
the sensitivity of surface waters to acidic deposition by revealing the impor-
tance of soil composition and hydrologic flowpath, in addition to geology, in
delineating sensitive regions.
The geologic composition of a region plays a dominant role in influencing
the chemistry and, therefore, sensitivity of surface waters to the effects of
acidic deposition. Bedrock geology formed the basis for a national map of
surface water sensitivity (Norton et al., 1982) and has been used in numer-
ous acidification studies of more limited extent (e.g., Bricker and Rice, 1989;
Dise, 1984; Gibson et al., 1983). Analysis of bedrock composition continues
to be an important element for assessing sensitivity of surface waters in
mountainous regions (e.g., Stauffer, 1990; Stauffer and Whittchen, 1991;
Vertucci and Eilers, 1993).
The presence of large populations of acidic and low-ANC lakes and
streams in regions such as Florida that are underlain by calcareous bedrock
illustrate that if the surface waters are isolated from highly weatherable bed-

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38

Aquatic Effects of Acidic Deposition

hydrology as a key component associated with changes in the acid-base chem-
istry of lakes included in EPA's Long Term Monitoring Program.

3.2 Causes of Acidification

3.2.1 Sulfur

Several watershed processes control the extent of ANC generation and its
contribution from soils to drainage waters as acidified water moves through
undisturbed terrestrial systems. These are the major processes that regulate
the extent to which drainage waters will be acidified in response to ambient
levels of acidic deposition. Of particular importance is the concentration of
acid anions in solution. Naturally occurring organic acid anions, produced in
upper soil horizons, normally precipitate out of solution as drainage water
percolates through lower mineral soil horizons. Soil acidification processes
reach an equilibrium with acid neutralization processes (e.g., weathering) at
some depth in the mineral soil (Turner et al., 1990). Drainage waters below
this depth generally have high ANC. The addition of strong acid anions from
atmospheric deposition allows the natural soil acidification and cation leach-
ing processes to occur at greater depths in the soil profile, thereby allowing
water rich in mobile anions such as SO

4


4
2-

through watersheds is only one part
of a complex set of watershed interactions that govern the response of both
aquatic and terrestrial ecosystems to acidic deposition.

3.2.2 Organic Acidity

Organic acids commonly exert a large influence on surface water acid–base
chemistry, particularly in dilute waters having moderate to high dissolved

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Chronic Acidification

39
organic carbon (DOC) concentrations. Some lakes and streams are naturally
acidic as a consequence of organic acids in solution. The presence of organic
acids also provides buffering to minimize pH change in response to changes
in the amount of mineral (e.g., SO

4
2-

, NO

3


M) often contain wetlands and/or extensive organic-rich riparian
areas (Hemond, 1990).
Specification of the acid–base character of water high in DOC is somewhat
uncertain. Attempts have been made to describe the acid–base behavior of
organic acids using a single H

+

dissociation constant (pK

a

), despite the fact
that organic acids in natural waters are made up of a complex mixture of
acidic functional groups. It has also been assumed in the past that organic
acids are essentially weak acids, whereas a portion (perhaps one-third) of the
acidity is actually quite strong, with some ionization occurring at pH values
well below 4.0 (Hemond, 1994; Driscoll et al., 1994). A number of modeling
approaches have been used to estimate the acidity of organic acids in fresh
waters, often as simple organic acid analogs having different pK

a

values
(Oliver et al., 1983; Perdue et al., 1984; Driscoll et al., 1994).
In lakes sampled by the ALSC, estimated values of organic acid anion
concentration per mol DOC (RCOO

-

-

/DOC at 0.1 pH unit intervals, as a function of pH for the reduced ALSC data
set included in the analyses of Driscoll et al. (C.T. Driscoll, M.D. Lehtinen, and T.J. Sullivan,
1994, Modeling the acid-base chemistry of organic solutes in Adirondack, NY, lakes,

Water
Resour. Res.

, Vol. 30, p. 301, Figure 1; copyright by the American Geophysical Union. With
permission.)

B
A

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Chronic Acidification

41
The ALSC data were fitted by Driscoll et al. (1994) to a triprotic organic acid
analog representation that provided a good fit to the data (

r

2

= 0.92), with pK


water pH and ANC in this large database (Driscoll et al., 1994).
The importance of naturally occurring organic acids as agents of surface
water acidification has recently been substantially reinforced by several
modeling studies (e.g., Sullivan et al., 1994, 1996). These have shown that
inclusion of organic acids in the MAGIC model have an appreciable effect
on model predictions of surface water pH, even in waters where DOC con-
centrations are not particularly high. Concern was raised subsequent to
NAPAP’s Integrated Assessment (NAPAP, 1991) regarding potential bias
from the failure to include organic acids in the MAGIC model formulations
used in the IA. MAGIC hindcasts of pre-industrial lake-water pH of
Adirondack lakes showed poor agreement with diatom inferences of pre-
industrial pH. Revised MAGIC simulations, therefore, were constructed
that included the organic acid analog model developed by Driscoll et al.
(1994). The revised MAGIC hindcasts of pre-industrial lake-water pH that
included an organic acid representation showed considerably closer agree-
ment with diatom inferences (Figure 3.3). The mean difference between
MAGIC and diatom estimates of pre-industrial pH was reduced from 0.6
pH units to 0.2 pH units when organic acids were included in the model,
and the agreement for individual lakes improved by up to a full pH unit
(Sullivan et al., 1996).
Inclusion of organic acids in the MAGIC simulations for watershed manip-
ulation data sets at Lake Skjervatjern (Norway), Bear Brook (Maine), and Ris-
dalsheia (Norway) also had dramatic effects on model simulations of pH. In
all cases, MAGIC simulated considerably higher pH values when organic
acids were omitted from the model. Even at Bear Brook, where annual aver-
age DOC concentrations are very low (less than 250

µ

M C), incorporation of

) as much as an exchange of SO

4
2-

and NO

3
-

anions for
organic anions, with little or no change in ANC and pH.
Data are scarce with which to directly evaluate the hypothesis that acidic
deposition causes decreased organic acidity, but a variety of indirect evidence
was summarized in the review of Marmorek et al. (1988). They concluded
that there were a number of inconsistencies in the available data, but most
data suggested that organic acids have been lost from lake water as a conse-
quence of acidic deposition. Hypothesized mechanisms included:

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Chronic Acidification

43

FIGURE 3.4

MAGIC simulated pH with and without inclusion of the triprotic organic acid analog, and
observed pH, in the treatment and control lake/stream at (A) Skjervatjern, (B) Risdalsheia, and

1985) and pH dependent changes in dissociation of organic acids (Oliver et
al., 1983; Wright et al., 1988b) appeared most likely to be significant. Quanti-
tative estimates of change in DOC were not possible, but based on the avail-
able data, Marmorek et al. (1988) concluded that potential DOC losses of up
to 250

µ

M C were not unreasonable. Subsequent research has suggested,
however, that decreases in DOC concentrations in surface water in response
to acidic deposition have probably been less than 250

µ

M C (Wright et al.,
1988b; Kingston and Birks, 1990; Cumming et al., 1992). Furthermore, Krug
and co-workers contended that interactions between acidic deposition and
organic matter can either increase or decrease DOC, depending on the nature
of the organic matter interacting with the acid (e.g., Krug et al., 1985; Krug,
1991a,b).
Kingston and Birks (1990) presented diatom-based paleolimnological
reconstructions of DOC for lakes studied in the Paleoecological Investigation
of Recent Lakewater Acidification (PIRLA-I) project. The DOC optima and
tolerances of diatom taxa in four regions (Adirondack Mountains, northern
New England, northern Great Lakes states, and northern Florida) were esti-
mated using maximum likelihood and weighted averaging regression. The
cumulative fit per taxon as a fraction of the taxon's total variance revealed
that few taxa were consistent in terms of their explanation of the DOC gradi-
ent from region to region. DOC explained a small, but significant, amount of
taxon variance in lakes in the Adirondack Mountains, northern Florida, and

In addition to potential changes in DOC concentrations in response to acidic
deposition, acidification or recovery can alter the charge density of organic sol-
utes and, thus, influence organic contributions to acidity (e.g., Wright et al.,
1988b). David et al. (in press) found that charge density of organic acids
decreased by about 1

µ

eq/L/mg C at West Bear Brook in response to 6 years of
experimental acidification, probably owing to greater protonation of organic
acid anions at the lower pH. There was no evidence of a change in DOC, how-
ever, in response to the acidification. Similar results were reported by Lydersen
et al. (1996) at Lake Skjervatjern in Norway. Loss of DOC in response to acidic
deposition can also cause a shift in Al species composition towards lesser com-
plexation with organic ligands. Such a shift from organic to inorganic Al
increases toxicity of the Al to aquatic biota (Baker and Schofield, 1982).
Hedin et al. (1990) artificially acidified a small, moderately high-DOC (725

µ

M C) stream with H

2

SO

4

at the Hubbard Brook Experimental Forest (HBEF)
in New Hampshire. The ambient stream-water pH (4.4) was near the range


µ

eq/L. Thus, the overall capacity of organic anions to neu-
tralize mineral acid inputs offset about 11% of the added H

2

SO

4

concentration
(Hedin et al., 1990). This experiment only considered interactions between
H

2

SO

4

and organic matter within the stream. Any additional buffering that
may have been provided within the terrestrial catchment was not repre-
sented in the experimental design. Also, any possible catchment-mediated
influences of the experimental acidification on organic acid properties, DOC
mobilization, and so on, were excluded from the experiment because the acid
was not applied to the catchment soils.
Webster et al. (1990) reported dramatic changes in lake-water ANC in Nev-
ins Lake, MI, in response to the effects of drought on the local hydrology.

2+

; DOC decreased slightly (approximately 50

µ

M C) and color
decreased by half in the acidified north basin after acidification to pH equal
to 4.7. Average DOC concentrations and color values of the 2 basins were sig-
nificantly different at pH 4.7, but not at higher pH values. In contrast, Schin-
dler and Turner (1982) did not find significant changes in lakewater DOC in
response to the artificial acidification of Lake 223 in the Experimental Lakes
Area of Ontario to pH approximately equal to 5.4.
Sullivan et al. (1994) examined the results of the three catchment manipu-
lation experiments (Bear Brook, Maine, and Lake Skjervatjern and Ris-
dalsheia, Norway) that were conducted by Norton et al. (1993), Gjessing
(1992), and Wright et al. (1993). All three catchments showed some evidence
of changes in organic acid anion concentration in response to experimental
acidification or de-acidification treatment. Changes in DOC also may have
occurred. Unfortunately, however, none of these manipulation experiments
provided conclusive quantitative data regarding the effects of acidification
on DOC mobilization from catchment to surface waters or changes in the
concentration of DOC caused by acidification. There were problems in inter-
pretation of the data regarding changes in the concentration of dissolved or
total organic C (DOC/TOC) in runoff from each of the studies.
In the Watershed Manipulation Project at Bear Brook, DOC declined
about 50% from 1989 to 1992 in both East (reference catchment) and West
(treatment catchment) Bear Brooks. These streams were very low in DOC
throughout most of the year (annual average DOC less than 300



47
At Risdalsheia, the annual variability in TOC was very large at both the
roofed control and manipulated catchments and sufficient pretreatment data
were not collected to allow a rigorous evaluation of the extent to which TOC
mobilization may have been affected by the acid exclusion. Thus, none of
these three experimental manipulation studies provide the kind of quantita-
tive data on DOC/TOC responses to acidification that would be needed to
justify modifying predictive models to account for hypothesized changes in
the concentration of organic C in response to changes in acidic deposition.
Based on results available to date, it appears that changes in DOC concen-
tration in response to changes in acidic deposition may occur but are gener-
ally small in magnitude. The concentration of organic acid anions is affected,
however, by changes in acidic deposition, particularly in high-DOC waters.
This change can be appreciable in some cases and organic acids can provide
significant buffering against pH change in watersheds that receive acidic
deposition. For example, results of a resurvey of 485 Norwegian lakes sam-
pled in both 1986 and 1995 provided evidence in support of an increase in
organic acid anion concentrations in association with decreased lake-water
SO

4
2-

concentration (Skjelvåle et al., 1998). On a regional basis, the organic
acid anion concentration increased by an amount equal to between 9 and
15% of the decrease in SO

4
2-


3.2.3 Nitrogen

Nitrate (and also NH

4
+

that can be converted to NO

3
-

within the watershed)
has the potential to acidify drainage waters and leach potentially toxic Al
from watershed soils. In most watersheds, however, N is limiting for plant
growth and, therefore, most N inputs are quickly incorporated into biomass
as organic N with little leaching of NO

3
-

into surface waters. A large amount
of research has been conducted in recent years on N processing mechanisms
and consequent forest effects, mainly in Europe (Sullivan, 1993). In addition,
a smaller N research effort has been directed at investigating effects of N dep-
osition on aquatic ecosystems. For the most part, measurements of N in lakes
and streams have been treated as outputs of terrestrial systems. However,
concern has been expressed regarding the role of NO


program, relatively little attention was paid to N research.
Concern for chronically elevated NO

3
-

concentrations in aquatic ecosys-
tems received considerably greater attention in 1988 following the publica-
tion of the resurvey of Norwegian lakes (SFT, 1987). Over 1000 lakes, 305 of
which were originally sampled in 1974/75 (Wright and Henriksen, 1978),
were sampled again in 1986 (SFT, 1987; Henriksen and Brakke, 1988). Even
though the average SO

4
2-

concentration declined in the lakes, the pH
remained virtually unchanged because of increased NO

3
-

and decreased base
cation concentrations. In the southern portions of Norway, NO

3
-

concentra-
tions in the lakes doubled between 1974–1975 and 1986, reaching county-

-

. Although SO

4
2-

remained the dominant anion in most systems, the ratio of NO

3
-

/(NO

3
-

+
SO

4
2-

) reached 0.54 on an equivalent basis in some lakes and rivers in south-
western Norway (Henriksen and Brakke, 1988). These authors summarized
the ratio of NO

3
-



concentration increase
between 1986 and 1995, and the average increase was only 1

µ

eq/L
(Skjelkvåle et al., 1998).
Increased atmospheric deposition of N does not necessarily cause adverse
environmental impacts. In most areas, added N is taken up by terrestrial
biota and the most significant effect is an increase in forest productivity
(Kauppi et al., 1992). However, in some areas, especially at high elevation
sites, terrestrial ecosystems have become N saturated* and high levels of dep-
osition cause elevated levels of NO

3
-

in drainage waters (Aber et al., 1989,
1998; Stoddard, 1994). This enhanced leaching of NO

3
-

causes depletion of

* The term nitrogen-saturated has been defined in a variety of ways, all reflecting a condition
whereby the input of nitrogen (e.g., as nitrate, ammonium) to the ecosystem exceeds the require-
ments of terrestrial biota and a substantial fraction of the incoming nitrogen leaches out of the
ecosystem in groundwater and surface water.

3
-

concentrations greater than 10

µ

eq/L. Of those, 5
lakes were situated at low elevation (less than 500 m) in the state of Washing-
ton and had relatively high ANC (greater than about 200

µ

eq/L). Because of
the high neutralization capacity, the N concentrations did not have a signifi-
cant impact on chronic acid–base status of these lakes. The other 19 high NO

3
-

lakes were all situated at high elevation, most above 3000 m. Cold tempera-
tures in such lakes undoubtedly play a major role in maintaining chronically
elevated NO

3
-

concentrations, largely by limiting biological uptake processes
in both the aquatic and terrestrial environments. The high NO



3
-

acidity which could be very
important biologically.
The Uinta Mountains of Utah and the Bighorn Mountains of central Wyo-
ming had the greatest percentages of high NO
3
-
lakes in the West, irrespec-
tive of lake-water ANC, with 19% of the lakes included within the Western
Lakes Survey having NO
3
more than 10 µeq/L. This is a high percentage of
lakes with measurable NO
3
-
for fall samples and indicates that NO
3
-
depo-
sition in these areas may have exceeded the capability of these systems to
assimilate N. It is unknown if these concentrations of NO
3
-
represent
impacts from anthropogenic sources or if this constitutes an unusual natu-
ral condition associated with inhibited NO
3

of phytoplankton in alpine lakes by factors other than N (e.g., P, temperature;
Baron et al., 1994). See further discussion of this topic in Chapters 7 and 11.
Recent research results regarding N dynamics and the effects of elevated N
deposition are considered in greater detail in Chapter 7.
3.2.4 Base Cation Depletion
Calcium and other base cations are important nutrients that are taken up
through plant roots in dissolved form. Base cations are typically found in
abundance in rocks and soils, but a large fraction of the base cation stores are
bound in mineral structures and are unavailable to plants. The pool of soluble
base cations resides in the soil as cations that are adsorbed to negatively
charged exchange sites. They can become desorbed in exchange for H
+
or Al
3+
and are, thus, termed exchangeable cations. The process of weathering grad-
ually breaks down rocks and minerals, returning their stored base cations to
the soil in dissolved form and, thereby, contributing to the pool of adsorbed
base cations. Base cation reserves are gradually leached from the soils in
drainage water, but are constantly being resupplied through weathering.
It has long been recognized that elevated leaching of base cations by
acidic deposition might deplete the soil of exchangeable bases faster than
they are resupplied via weathering (Cowling and Dochinger, 1980). How-
ever, base cation depletion of soils had not been demonstrated at the time
of the Integrated Assessment. Scientific appreciation of the importance of
this response has increased with the realization that watersheds are gener-
ally not exhibiting ANC and pH recovery in response to recent decreases in
S deposition. In many areas, this lack of recovery can be at least partially
attributed to decreased base cation concentrations in surface water. The
base cation response is quantitatively more important than was generally
recognized in the 1980s.

believed. This understanding developed in large part from paleoecological
studies (e.g., Davis et al., 1988; Charles et al., 1990; Sullivan et al., 1990a) that
concluded that historical changes in lake-water pH and ANC were small rel-
ative to estimated increases in lake-water SO
4
2-
concentrations since pre-
industrial times.
After passage of the Clean Air Act in 1970 and subsequent amendments in
1990, emissions and deposition of S were reduced and the concentrations of
SO
4
2-
in lake and stream-water in the eastern U.S. and Canada decreased (Dil-
lon et al., 1987; Driscoll et al., 1989a; Sisterson et al., 1990). Long-term moni-
toring data confirmed that much of the decrease in surface water SO
4
2-
concentration was accompanied by rather small pH and ANC recoveries
(Driscoll and van Dreason, 1993; Kahl et al., 1993b; Driscoll et al., 1995; Likens
et al., 1996). The most significant response, on a quantitative basis, was
decreased concentrations of Ca
2+
and other base cations. Similarly, long-term
monitoring data from four small watersheds in Norway illustrated substan-
tial declines in both S deposition and stream-water SO
4
2-
concentration since
the late 1970s. Reductions in SO

reduced emissions and deposition of S and N.
Thus, as SO
4
2-
concentrations in lakes and streams have declined so, too,
have the concentrations of Ca
2+
and other base cations. There are several
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52 Aquatic Effects of Acidic Deposition
apparent reasons for this. First, the atmospheric deposition of base cations
has decreased in recent decades (Hedin et al., 1994), likely owing to a combi-
nation of air pollution controls, changing agricultural practices, and the pav-
ing of roads (the latter two affect generation of dust that is rich in base
cations). It has been estimated that more than half of the supply of Ca
2+
to cat-
ion pools of forest soils in the northeastern U.S. may be derived from atmo-
spheric inputs. Similarly, Driscoll et al. (1989a) estimated that between 77 and
85% of the decline in the concentration of base cations in stream water at
Hubbard Brook Experimental Forest (HBEF) could be attributed to decreased
base cation deposition. However, atmospheric deposition of base cations has
increased in Maine since 1982 concurrent with large declines in the concen-
tration of base cations in drainage lakes. Moreover, the concentrations of base
cations in groundwater recharge seepage lakes have not declined, which sug-
gests that watershed processes have been altered and that changes in base
cation deposition are not responsible for changes in the concentration of base
cations in lake waters in Maine (Kahl, personal communication). Second,
decreased movement of SO

to Al
n+
ratio in soil solution (Cronan
and Grigal, 1995).
Lawrence et al. (in press) investigated base cation dynamics in soils in the
Neversink River Basin in the Catskill Mountains, NY. They found that S dep-
osition increased along an elevational gradient, whereas the concentrations
of soil exchangeable bases decreased with elevation. A large quantity of soil
was collected from a low-elevation site, bulked, and then redistributed to
about 30 sites along the elevational gradient. At each site, soil was placed in
mesh bags, buried, and then retrieved and analyzed after 1 year. Results of
chemical analyses confirmed that the concentration of exchangeable bases
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Chronic Acidification 53
and the base saturation* decreased with increasing elevation (Lawrence et al.,
in press). Field data and laboratory analyses of soil samples were consistent
with the interpretation that observed decreases in ANC of stream water in the
Neversink River watershed since 1984 have been the result of decreased base
saturation of soils caused by acidic deposition (Lawrence et al., in press).
Lawrence et al. (1995) proposed that the dissolution of Al in the mineral soil
by mineral acid anions supplied by acidic deposition (SO
4
2-
, NO
3
-
) can
decrease the availability of Ca
2+

deposition of S or N, exposure to ozone, natural disturbance factors such as
wind and fire, and climatic changes.
Landscape processes affect the acid–base chemistry of drainage waters in a
variety of ways. Some processes contribute to the acidification of soil and sur-
face waters or reduce the base saturation of the soils thereby increasing their
sensitivity to acidic deposition. Other processes cause decreased acidity (Sul-
livan et al., 1996b; Table 3.1).
Disturbances such as logging, blowdown, and fire affect surface water pH
and ANC. Watershed disturbance disrupts the normal flow of water, in some
cases causes increased contact between runoff water and soil surfaces, and
often leads to increased base cation concentration and ANC in surface
waters. Recovery from disturbance will, in most cases, lead to a decrease in
*

Base saturation is the concentration of exchangeable base cations as a percentage of the total
cation exchange capacity, which also includes H
+
and Al
n+
.
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© 2000 by CRC Press LLC
54 Aquatic Effects of Acidic Deposition
pH and ANC as the system returns to predisturbance conditions. A short-
term investigation of an ecosystem in the process of recovering from a water-
shed disturbance might erroneously conclude that acidification was occur-
ring in response to changes in atmospheric deposition or some other cause
external to the watershed.
The influence of historical forest management on the ability of a given for-
est ecosystem to process N is not well understood. Nevertheless, forest man-

deposition inputs
Road building and
construction
More base cation neutralization, less acidity initially
Depletion of base cation reserves in soils, more acidity long term
Drainage of
wetlands
Re-oxidation of stored sulfur, pulses of acidity with increased
discharge
Drought Reduced groundwater inputs to seepage lakes with consequent
increased acidity
Increased relative baseflow to drainage waters with consequent
decreased acidity
Lake shore
development
Decreased acidity
Insect damage Pulse of nitrate acidity initially
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Chronic Acidification 55
where land use changes presumably have not occurred. Whereas it is true that
landscape change often occurs as a manifestation of land use change, equally
dramatic landscape changes can occur in response to natural factors without
change in the way that humans use the land. Also, the documented occurrence
of acidification in the absence of either land use or landscape change does not
negate the importance of other questions concerning the interactions between
such changes and acidic deposition (Sullivan et al., 1996b).
Land management activities, particularly removal of forest or change in
forest structure, have important effects on hydrology and the total deposition
of S, N, and marine salts. An important land use change during the last 60

drainage water quality through vegetation uptake processes. This is because
trees accumulate base cations to a greater degree than anions. In order to bal-
ance the resulting charge discrepancy, roots release an equivalent amount of
protons. This is an acidifying process. Base cation accumulation by growing
trees is strongly age dependent. Young, fast-growing forests are more acidi-
fying than older forests (Nilsson et al., 1982; Nilsson, 1993) and retain greater
amounts of N inputs. For example, Reynolds et al. (1994) found concentra-
tions of NO
3
-
in 136 streams in upland Wales were significantly correlated (p
< 0.001) with the average age of conifers.
It has been proposed that forest blowdown affects surface water acid–base
chemistry via changes in hydrologic flow (Dobson et al., 1990). Pipes formed
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56 Aquatic Effects of Acidic Deposition
in the soil by tree roots can alter hydrologic flow so that less water enters the
soil matrix, where neutralization processes buffer the acidity of incoming
rainwater and snowmelt. Pipes tend to be located in near-surface soil hori-
zons where most tree rooting occurs, and contact between drainage waters
and mineral soil is reduced when runoff is routed through them. If enhanced
pipeflow affects a large portion of any watershed, stream and lake chemistry
may be expected to reflect the chemical characteristics of surface and near-
surface soil waters more so than the characteristics of deeper groundwater
and more so than would be the case in the absence of such pipeflow.
During the 1980s, the prevailing scientific consensus held that the majority
of lakes in eastern North America that had pH less than about 5.5 to 6.0 had
been acidified by acidic deposition. Temporal/spatial correlations were
claimed to support this contention. Reports that acidic surface waters were

As discussed previously, land use changes and disturbances within the
drainage basins of lakes and streams can influence water chemistry, but the
regional acidification of surface waters in parts of Europe and North America
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© 2000 by CRC Press LLC
Chronic Acidification 57
has not been attributed to changes in land use practices. In many cases, such
disturbances increase ANC and pH, and cause water quality problems other
than acidification. Where land use changes have been substantial, it may be
difficult to quantify the effects of acidic deposition on a regional scale. A crit-
ical limitation of much of the acidic deposition effects research conducted to
date has been, however, the nature of the questions being asked. In the major-
ity of cases, land use has been addressed only as a potential alternative expla-
nation for acidification, rather than being evaluated in an open and objective
fashion (e.g., Havas et al., 1984; Birks et al., 1990b).
There has not been a rigorous regional evaluation of land use changes in
areas of the U.S. susceptible to acidic deposition effects. In the absence of
such an investigation, it has not been possible to quantify the extent or mag-
nitude of land use related effects on water quality within the regions of con-
cern. It is clear, however, that such changes can have important effects on
acid–base status.
Renberg et al. (1993) evaluated sediment composition, pollen, radiocarbon
and lead dating, and diatom reconstructions to ascertain the effects of chang-
ing land use in 14 widely separated lakes in southern Sweden over the past
10,000 years. Lake-water pH declined from about 7.0 to 5.5 in the first few
thousand years after deglaciation in response to natural processes. During
the Iron Age, the area was deforested as an agrarian economy developed in
the region: this caused an increase in lake-water pH of 0.5 to 1.4 pH units in
12 of the 14 study lakes. Subsequently, pH declined during the nineteenth
and twentieth centuries, following abandonment of agriculture, to levels less

vated S or N deposition in several ways. Drought can alter hydrologic flow-
paths and change the relative contribution of near-surface runoff vs. deeper
baseflow. Because these source areas typically generate different levels of
ANC, such changes in hydrologic input can profoundly influence surface
water acid–base chemistry (Webster et al., 1993; Newell, 1993).
The volume of annual precipitation received by a watershed, especially
during winter, has been shown to dramatically affect the total annual wet
deposition of S and N to that watershed. Because a relatively large proportion
of the snowpack ionic load is released during the early phases of snowmelt,
high-elevation western watersheds are potentially exposed to greater epi-
sodic acidification during years with greater precipitation.
Climate warming can influence the response of surface waters to past and
future acidic deposition. Under cool, moist conditions, a sizable component
of the atmospheric S inputs can be stored as reduced S in soils, especially in
wetland areas. This storage protects surface waters from acidification (Roch-
efort et al., 1990). However, under warmer and drier climatic conditions, this
stored S can be reoxidized and, consequently, released to drainage waters
during periods of rainfall or snowmelt (Bayley et al., 1992; LaZerte, 1993;
Schindler, 1998).
Temperature can also have a variety of effects on S and N dynamics. The
timing and rapidity of snowmelt are important factors governing the deliv-
ery of ionic loads from the snowpack to surface waters. Temperature also
has a large influence on biological uptake of N within both terrestrial and
aquatic ecosystems.
Drought conditions in the Sierra Nevada were judged by Melack et al.
(1998) to be responsible for increasing the proportion of runoff derived from
shallow groundwater in the Ruby Lake basin, as evidenced by an increase in
SO
4
2-

© 2000 by CRC Press LLC
Chronic Acidification 59
Chorover et al. (1994) evaluated the effects of fire on soil and stream-water
chemistry in Sequoia National Park. Burning increased concentrations of
NO
3
-
and SO
4
2-
in soils and stream water. Sulfate increased 100 fold, and NO
3
-
remained higher in soils and stream water for 3 years. Fenn and Poth (1998)
hypothesized that successful fire suppression efforts may have contributed to
the development of N saturation in fire-adapted ecosystems in southern Cal-
ifornia by allowing N to accumulate in soil and in the forest floor, and by
maintaining dense overmature stands with reduced N demand.
3.2.8 Hydrology
The impacts of atmospheric deposition on high-elevation aquatic systems are
strongly controlled by the flowpaths of water through the catchments.
Hydrology is an important controlling factor for deposition impacts in virtu-
ally all environments (Turner et al., 1990), but hydrology is of overriding
importance in alpine and subalpine ecosystems, such as are found through-
out the West. The depth and make-up of soils, talus, and colluvium, and the
slope of the watershed collectively determine the residence time of subsur-
face water within the watershed, the extent to which snowmelt and rainfall
runoff interact with soils and geologic materials, and consequently the extent
of NO
3

infiltrated the soil and talus areas. After the pre-event soil waters have been
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