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3-98
30
20
10
n
RED CHALK No. 4
-
^^\
_l 1 1 1 1 1 1 L. __L. ±....1111 II 1 II II
70 r
60
50
40
30
20
10
RED CHALK No. 3
1900 2100 2300 0100 0300 0500 0700 0900 1100 1300 1500
July 12 / 77
Rain Event
Time (hr)
July 13 177
pH 4.06 ~2cm
Figure 3-36.
Hydrogen ion content of streams draining Red Chalk Lake
watersheds No.3 and No.4 (Muskoka-Haliburton, Ontario)
showing effects of a 2 cm rainfall (pH 4.06) between
11:00 p.m. July 12, 1977 and 3:00 a.m. July 13, 1977
(Scheider et al. 1979b).
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3-99
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-------
3-100
3.7 ALTERATION OF BIOTIC COMPONENTS IN AQUATIC SYSTEMS RECEIVING
ACIDIC DEPOSITION
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U.S. Forest Service in Monongahela National Forest (Dunshie 1979).
Due to the sandstone geology of the watershed, the tributaries and
the river are poorly buffered and subject to rapid changes in water •
quality. The lowest pH values in both streams (Little Black Fork is •
a control area, with no logging or coal mining) normally occurred
during the winter and early spring, apparently because of snowpack IB
melting. The highest pH occurred during low stream flow periods in jf
the summer and fall. Even though summer and autumn are the periods
of highest precipitation inputs (see below), more extensive contact H
between soils and precipitation may have lead to greater neutrali- •
zation at these times than during either winter or spring flushing
events.
The effect of rainfall on river pH is more apparent when individual «•
events are examined. A graphic presentation pairing daily river and
precipitation events with pH during summer periods is shown in •
Figure 3-37. During the growing season, a storm event with a j§
subsequent increase in discharge can significantly lower river pH
below the natural nonstorm daily variation. The magnitude of this •
downward shift is dependent upon rainfall characteristics (pH, •
amount, intensity, and area distribution) and antecedent soil
moisture. Downward shifts in river pH ranging from 0.6 to 0.9 pH
units, occurred on July 11 and 26, and on August 15 and 25, 1977. On •
three of these days, at least 3.3 cm of rainfall fell within a •
48-hour period; pH of the rainfall for these dates ranged from 3.7 to
4.2. •
Nearly 13 years of pH data have been collected at the Bowden Fish
Hatchery river intake on the Shavers Fork River, showing lower pH .
values during winter and spring compared to summer conditions. This •
is important for aquatic organisms and has been measured in other
poorly buffered streams. This pH trend occurred in streams and
tributaries independent of watershed disturbance by mining (Dunshie •
1979). I
I
Many changes in biota have been linked to acidification of surface •
waters. In some controlled whole lake and laboratory experiments a *
causal relationship with decreased pH has been established. In the
majority of cases, the observed changes in biota have simply been fl
correlated with observed changes in pH and other parameters, but |
causality has not been established. For many biological communities,
acidification has been accompanied by decreases in species diversity •
and changes in species dominance. Acidification may also be •
accompanied by species extinctions, or decreases in overall community
standing stocks. This topic has been reviewed recently by Raines
(1981c). Generalized summaries of responses of aquatic organisms to •
low pH are given in Figures 3-38 and 3-39 (Eilers and Berg 1982), and ™
are presented here as a simplified overview of the complex
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3-101
A Rainfall pH
I Rainfall Accumulation
10 15 20 25 30
5 10 15
7/77
Figure 3-37.
Mean daily pH for the Shavers Fork River at Bemis, West
Virginia and precipitation event pH and accumulation at
Arborvale, West Virginia (Dunshie 1979).
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3-102
Algae
Insects
Molluscs
Sponges
Leeches
Zooplankton
Fish
Frogs
75
50
25
75-
50
25
75
50-
25
75
50
25
75
50
25
75
50
25
75
50
25
PH
Figure 3-38. Relative number of taxa of the major taxonomic groups
as a function of pH (Eilers and Berg 1982).
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3-103
c
0
O
L_
CD
a.
100
75 -
50 -
25 -
0
Major Aquatic
Community Impacts
yMi i n*lit*i i I f-
1 I 3 snails5
zooplankton I insects I I phytoplankton
CO
•^
CO
1 Sprules (1975) - Ontario
2 Beamish (1976) - Ontario
3 Bell (1971) - Laboratory TL50
4 Yan and Stokes (1978) -
- Ontario
PH
5 0kland (1969) - Scandinavia
6 Wright et al. (1976) - Norway
7 Kwiatkowski and Roff
(1976) - Ontario
8 Snekvik (1974) - Norway
Figure 3-39. Generalized response of aquatic organisms to low pH
(Eilers and Berg 1982).
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3-104
3.7.1 Effects on Algae
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interactions described below. It is important to note that the data
were derived from literature surveys of the relationship between the
distribution of groups of organisms versus lake and stream pH values. •
The quantitative description of these relationships may not reflect •
the response of individual taxa.
Definitive experiments are required to demonstrate whether such •
changes are directly attributable to increases in hydrogen ion •
concentration or whether they are attributable to secondary ecosystem
interactions, such as elevation of trace metal levels or disruptions B
of normal food chains. In spite of incomplete understanding of the |
actual mechanisms underlying observed changes accompanying pH
declines, it appears that acidification of surface waters brings •
about major quantitative and qualitative changes in structure and •
function of aquatic ecosystems. Disruption of the normal food chains
may occur long before the lakes have been acidified in a chemical
sense. •
I
The free-floating (planktonic) and attached (benthic and epiphytic)
algae are the major primary producers in most aquatic ecosystems and «
directly or indirectly provide most of the food for zooplankton and •
ultimately for fish. Evidence gathered mainly from synoptic surveys
in Scandinavia, Canada and the United States has indicated that the
species diversity of benthic and planktonic algal communities is less •
in acidified lakes. Yan and Stokes (1976) observed only nine species •
of phytoplankton in a single sample from Lumsden Lake (pH 4.4;
Beamish and Harvey 1972), in the La Cloche Mountains in Ontario, but •
observed over 50 species in each of two nearby nonacidic lakes, |
having pH over 6.0. Diversity indices for phytoplankton populations
in the La Cloche Mountain lakes are much less in lakes with pH values •
below 5.6 (Kwiatkowski and Roff 1976). In Scandinavian lakes numbers •
of phytoplankton species are also much less in lakes with pH values
below 5.5 (Aimer et al. 1978; Leivestad et al. 1976).
Some long-term functional adaptations to certain acidic environments •
may occur. Raddum et al. (1980) have suggested that such a mechanism
explains the observation that a group of relatively recently •
acidified clearwater lakes in Norway have less diverse phytoplankton Jj
assemblages than naturally acidic, humic lakes. Additionally, the
bioavailability and toxicity of trace metals may be lower in the «
brownwater acidic lakes because metals may be complexed with humic •
materials.
Although species diversity of phytoplankton generally decreases with •
increasing acidity, biomass (Yan 1979) and productivity (Aimer •
et al. 1978; Schindler 1980) are often not reduced by acidification.
However, if phosphorus (the nutrient that normally limits phyto- •
plankton productivity in soft-water lakes) is immobilized to some •
degree in acidic lakes because of complexation with aluminum and
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3-105
humic material (Aimer et al. 1978), this would result in reduced
primary productivity. To date, data from lakes in Scandinavia and
eastern Canada indicate no significant correlations between pH and
phytoplankton biomass or productivity (Harvey et al. 1981).
Phytoplankton communities of nonacidic oligotrophic lakes in eastern
Canada are typically dominated by chrysophytes (Schindler and
Holmgren 1971) or by diatoms (Duthie and Ostrofsky 1974). In
contrast, strongly acidic lakes are generally dominated by dino-
flagellates. In Sweden, the dinoflagellates, formed 85% of the
biomass in lakes of pH 4.6-5.5 (Dickson et al. 1975). Of 14 lakes in
central Ontario, dinoflagellates formed between 30 and 70% of the
phytoplankton biomass in 4 lakes having pH 4.2-4.8, but only 2-30% of
the biomass in 10 lakes with pH levels of 5.8-6.8 (Yan 1979).
In certain poorly buffered lakes, some of the phytoplankton species
may interfere with recreational use of the lakes. For example, in
five lakes in Ontario and New Hampshire with pH 5.5-6.2, obnoxious
odours developed during the summers of 1978, 1979, and 1980 (Nicholls
et al. 1981). The odours have been shown to be caused by the growth
of the planktonic Chrysochromulina breviturrita. This species was
first discovered in 1976, but it is now known to inhabit more than
40 lakes in Ontario, most of which are acidic (Nicholls et al. 1981).
The "invasion", and associated odour production, by this organism is
apparently a recent phenomenon. Although the relationship between
lake acidification and the proliferation of this species has not been
proven, data collected thus far indicate that dominance of this
species to an extent causing the serious odour production, is
restricted to acidic lakes.
Acidified lakes and streams are often characterized by increased
growth of benthic filamentous algae. In Sweden, Ontario and Quebec,
unusually dense and extensive masses of filamentous algae (mainly
Mougeotia, Zygogonium and Zygnema sp.) proliferate in the littoral
zones of many lakes with pH values of 4.5-5.5 (Blomme 1982; Grahn
et al. 1974; Hendrey et al. 1976; Hultberg and Grahn 1975; Schindler
1980; Stokes 1981). These filamentous algal growths are associated
either with macrophytes or other substrates or exist as floating
"clouds" near the lake bottom. The accumulations of algae may reduce
light availability to macrophytes, change microclimates for benthic
macroinvertebrates and restrict fish feeding and spawning. Some
depreciation of shoreline recreational values and activities,
especially swimming, may result from this growth of algae.
3.7.2 Effects on Aquatic Macrophytes
Information on the effects of acidification on macrophyte communities
of soft-water lakes is still incomplete. Scandinavian investigators
have suggested that when lake water pH declines, typical macrophyte
dominants are replaced by very dense beds of Sphagnum (Grahn et al.
1974; Hendrey et al. 1976; Hultberg and Grahn 1975). The loss of
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3-106
3.7.3 Effects on Zooplankton
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some macrophyte species and the correlative increase in Sphagnum
abundance may be indirectly related to depressed pH, through changes
in inorganic carbon availability (Raven 1970; Steemann-Nielsen 1944, I
1946). In Scandinavia, the decline of macrophyte species and the •
concurrent Sphagnum invasion begins as pH falls to about 6.0, and
proceeds rapidly when pH falls below 5.0. In Lake Golden in New York •
(pH 4.9), Sphagnum is abundant (Hendrey and Vertucci 1980), and in •
Beaverskin Lake in Kejimkujik National Park in Nova Scotia, a clear
lake of pH 5.3, Kerekes (1981) has reported extensive Sphagnum _
growth. In Ontario lakes, some species of Sphagnum have been I
identified (Harvey et al. 1981), but accumulations as dense as those ™
recorded in Scandinavia have not been observed.
I
Sphagnum moss coverage of littoral zones creates a unique habitat
that is considered unsuitable for some species of benthic inverte-
brates or for use as fish spawning and nursery ground (Hultberg and •
Grahn 1975). It may reduce the appeal of freshwater systems for •
certain recreational activities. Through the release of hydrogen
ions and polyuronic acids, Sphagnum could acidify their immediate _
surroundings should they accumulate (Clymo 1963; Crum 1976). •
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Four major groups of animals contribute to zooplankton communities:
protozoans, rotifers, crustaceans and insects. Zooplankton are an m
important food for many species of fish, particularly for younger •
individuals. Thus, they are an essential component of the aquatic
food chain, transferring energy and materials from the primary
producers (algae) to consumers, including fish and man. Acidifi- •
cation apparently results in reduced zooplankton biomasses, as both •
the numbers and average size of community members are reduced (Yan
and Strus 1980). As a result, food availability to higher trophic
levels may be decreased.
Acidification of lakes is accompanied by changes in the occurrence, «
abundance and seasonal succession of species, and in the diversity of •
crustacean (and other) zooplankton. It is often assumed that the
direct cause of these changes is differences in tolerance among
zooplankton species to increased H+ concentration. However, •
acidification also increases the transparency of lakes, increases the 9
concentration of potential toxicants such as Cd^+ (Aimer et al. 1978)
which is toxic to zooplankton at less than 1 Pg/L (Marshall and •
Mellinger 1980), and produces quantitative and qualitative changes in |
zooplankton predator and prey species (Harvey et al. 1981). Hence,
the immediate causes for the changes in zooplankton communities that _
do occur, while linked to increased acidity, may be quite complex. •
The most important components of zooplankton communities are usually
the rotifers and crustaceans. Of these, the crustaceans usually form I
90% of the biomass (Pederson et al. 1976), while rotifers, because I
they have shorter generation times, may be responsible for 50% of the
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3-107
zooplankton productivity (Makarawicz and Likens 1979). Available
studies on the effects of acidification on rotifer populations are
contradictory; both smaller (Roff and Kwiatkowski 1977) and larger
(Malley et al. 1982; Yan and Miller 1981) standing stocks have been
observed in acidic lakes. Studies from very acidic lakes (Smith and
Frey 1971) and the Smoking Hills Lakes of the Northwest Territories
(Havas 1980) indicate, however, that some species of rotifers can
survive when all crustacean zooplankton have been eliminated or could
not survive at the low pH conditions.
The diversity of zooplankton communities has been reported in several
studies to be greatly reduced by acidification (Raddum et al. 1980;
Sprules 1975). Whereas nonacidic lakes typically contain approxi-
mately ten species of planktonic Crustacea in mid-summer collections,
Sprules (1975) observed that the number of species and species
diversity of acidic lakes in the La Cloche Mountains in Ontario was
drastically reduced. In several cases only a single species,
Diaptomus minutus, remained.
The diversity of littoral cladocerans has also declined with
acidification (Brakke et al. 1982). The decrease in number of
species and diversity is apparently related to low pH and not to
changes in aquatic macrophytes (Kenlan et al. 1982). Sediment core
studies in New England and in Norway suggest that changes in littoral
cladoceran assemblages occurred simultaneously with calculated dates
of pH declines based on diatom analyses (Brakke et al. 1982; Davis
et al. 1982).
Some predacious zooplankton, for example cyclopoid copepods (Raddum
et al. 1980) and Epischura lacustris (Malley et al. 1982), are very
sensitive to acidification, and are often absent from acidic lakes.
Densities of other predators, such as some species of Chaoborus
(Eriksson et al. 1980a) and Heterocope saliens (Raddum et al. 1980),
apparently increase. The significance of these changes in predator
populations to zooplankton community structure is not yet understood
although it may be important (Eriksson et al. 1980a).
3.7.4 Effects on Aquatic Macroinvertebrates
Numerous aquatic macroinvertebrates are known to be affected by low
pH conditions. In some cases an entire phylum appears to be
affected, while in other situations susceptibility is species-
specific. Evidence indicates that molluscs, in general, are highly
susceptible to reduced pH (J. 0kland 1980; Raddum 1980; Wiederholm
and Eriksson 1977), often being restricted to habitats with pH
greater than 5.8-6.0. Similarly, all species of oligochaetes studied
thus far have been found at lower densities in acid waters
(Wiederholm and Eriksson 1977).
Sensitivity to low pH has been inferred from field investigations for
certain Arachnids, Crustaceans and Insects. Arachnids were only
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3-108
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briefly mentioned by Grahn and his co-workers (1974); acarinids were
absent in waters with pH values below 4.6. No macro-crustaceans were
found below pH 4.6 (Grahn et al. 1974). Gammarus lacustris was •
absent from waters with pH below 6.0 (J. j&kland 1969), while the |
crayfish, Astacus astacus was rare in lakes where the summer pH value
was less than 6.0 (Svardson 1974). Orders of Insecta exhibit a wide •
range of sensitivities to pH. While the numbers of species of •
Ephemeroptera and Plecoptera appear to be positively correlated with
pH, larvae of Chironimidae (Diptera), Hemiptera and Megaloptera are
often abundant in acid lakes (Aimer et al. 1978). Hutchinson et al. •
(1978) reported an example of extreme tolerance by larvae of red •
chironomids, Chironomus riparius, to waters of pH 2.2 in the
Northwest Territories.
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Although the field studies mentioned above provide evidence of the
effects of acidification on certain species, the pH of a natural &
system has rarely been altered experimentally, and the impacts on •
invertebrates noted. The documented effects of decreased pH include
the disappearance of Mysis relicta in Lake 223, an experimentally
acidified lake in the Experimental Lakes Area (Malley et al. 1982), •
elimination or reduction of Ephemeroptera populations in a stream in •
the Hubbard Brook Experimental Forest in New Hampshire (Fiance 1978;
Hall et al. 1980), and decreased emergence of some species of •
Plecoptera, Trichoptera and Diptera in the same stream (Hall et al. •
1980). Those species with acid-sensitive life stages (such as
emergence in insects) which can coincide with low pH snowmelt, or ^
other events, such as low pH flushing, may be especially sensitive to •
acid deposition. ™
In considering the distribution of the above species in relation to B
waters of varying pH no causative relationship between hydrogen ion •
concentration and the observed changes has been determined as yet.
Other factors vary with pH, including concentrations and availability •
of nutrients, bicarbonate, and various metals. From the results •
available, however, it appears that molluscs (perhaps because of
their requirement for calcium) and moulting crustaceans (perhaps _
because of their large demand for calcium at the time of moult) are •
the macroinvertebrates most sensitive to low pH levels. It is still •
unclear why certain groups of aquatic insects are more sensitive than
others. H
3.7.5 Effects on Bacteria and Fungi •
The decomposition rate of fixed carbon, both allochthonous and
autochthonous organic matter, is largely determined by microbial
processes in the water column and in the surface layers of sediment. •
Several studies have demonstrated that rates of decomposition of •
organic matter are decreased at low pH values. In a laboratory
study, for example, Bick and Drews (1973) demonstrated that as pH was
lowered, the number of bacteria and protozoans decreased, populations
of fungi increased, and the rates of decomposition and nitrification
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3-109
were reduced. Traaen and Laake (1980) measured decomposition rates
of homogenized birch litter and glucose/glutamate mixtures. When the
pH was decreased from 7.0 to 3.5, litter decomposition dropped to 30%
of control levels, and a shift from bacterial to fungal dominance was
observed. Traaen (1980) further observed that rates of weight loss
of decomposing birch leaves and aspen sticks after one year in the
laboratory or one to two years in field situations were significantly
lower at pH levels less than 5.0.
Reductions in numbers of heterotrophic bacteria have been observed
previously in aquatic habitats acidified by acid mine drainage
(Guthrie et al. 1978; Thompson and Wilson 1975; Tuttle et al. 1968,
1969). Caution must be exercised, however, in extrapolating results
from such studies to situations where the source of protons is
atmospheric because the pH is often much lower in acid mine drainage
lakes, and the concentration of dissolved substances, including
metals, much higher.
Rao et al. (1982) studied the effects of acidic precipitation on
bacterial populations of the Turkey Lakes, Ontario and Kejimkujik,
Nova Scotia. They observed reduced numbers of nitrifying bacteria
and sulphur cycle bacteria in low pH lakes and streams. Bacterial
activity as measured by oxygen consumption rate and biodegradation or
organic material was 50% less and 30-40% less respectively in
acid-stressed environments compared to nonacid-stressed areas.
Microbial transformations of sulphur and nitrogen species may
influence lake acidity and alkalinity (Brewer and Goldman 1976).
Schindler (1980) showed that increases in SO^- concentrations
stimulated sulphate-reducing bacteria in lakes that develop anoxic
hypolimnia. The reduction of SO^" yields OH~ thereby increasing
akalinity. Stimulation of SO/^- reduction has been used with success
to reclaim acid mine drainage waters. Sulphate-reducing bacteria
require anoxic conditions, and are stimulated by large quantities of
organic matter (i.e., they prefer conditions typical of eutrophic
lakes). However, acidified lakes are not eutrophic and many have
oxygenated hypolimnia.
3.7.6 Effects on Amphibians
Many species of frogs, toads, and salamanders breed in temporary
pools. These pools are formed by a mixture of snowmelt water and
spring rains and may have low pH values during the spring. Because
of the vulnerability of this habitat to pH depressions, amphibian
populations are expected to be one of the earliest forms of wildlife
to be affected by the acidification of fresh waters. Temporary pools
used as breeding sites by Jefferson's (Ambystoma jeffersonianum) and
yellow-spotted salamanders (A. maculatum) in New York were found to
have pH values 1.5 units lower than nearby permanent ponds (Pough and
Wilson 1977). The amphibian species of eastern Canada considered
most susceptible to the effects of acid deposition because of their
breeding habitat are listed in Table 3-19.
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3-110
TABLE 3-19. SUSCEPTIBILITY OF BREEDING HABITAT TO pH DEPRESSION
FOR THOSE AMPHIBIANS IN NORTHEASTERN NORTH AMERICA WHOSE RANGE
OVERLAPS AREAS RECEIVING ACIDIC DEPOSITION (modified from
Clark and Fischer 1981)
Potential for
acidification
of egg-laying
habitat Habitat Species
high
meItwater
pools
moderate
permanent
ponds
low
streams
lakes
bogs
logs and
stumps
Ambystoma maculatum - Yellow-spotted
salamander
Ambystoma laterale - Blue-spotted
salamander
Ambystoma tremblayi - Tremblays
salamander
Bufo americanus - American toad
Pseudacris triseriata - Chorus frog
Rana sylvatica - Wood frog
Rana pipiens - Northern leopard frog
Hyla crucifer - Northern spring peeper
Hyla versicolor - Gray tree frog
Necturus maculosus - Mudpuppy
Notophthalmus viridescens - Red-spotted
newt
Bufo americanus - American toad
Hyla versicolor - Gray tree frog
Pseudacris triseriata - Chorus frog
Rana catesbeiana - Bullfrog
Rana clamitans - Green frog
Rana pipiens - Northern leopard frog
Rana septentrionalis - Mink frog
Eurycea bislineata - Northern two-lined
salamander
Necturus maculosus - Mudpuppy
Rana catesbeiana - Bullfrog
Hemidactylium scutatum - Four-toed
salamander
Plethedon cinereus - Red-backed
salamander
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3-111
Detrimental effects of acidity on adult amphibians have been shown in
a number of field surveys. In England, Cooke and Frazer (1976)
reported that no adult newts were caught from ponds of pH less than
3.8. The natterjack toad (Bufo calamita) was not found in ponds
below pH 5 (Beebee and Griffin 1977) in England. The common toad
(Bufo bufo) did not occur where pH was less than 4.2, and the smooth
newt (Triturus vulgaris) occurred only rarely in ponds at pH values
less than 6.0. Hagstrom (1977) observed that the common toad and
common frog (Rana temporaria) disappeared when pH levels reached
4.0-4.5. In New Hampshire, when a section of Hubbard Brook was
artificially acidified to mean pH 4.0, salamanders disappeared from
the study area (Hall and Likens 1980).
Pough (1976) noted heavy embryonic mortalities and deformities in the
yellow-spotted salamanders which breed in temporary meltwater ponds
with pH less than 6.0. In central Ontario, Clark and Euler (1981)
reported that the numbers of egg masses of yellow-spotted salamanders
and male calling densities (an estimate of population size) of spring
peepers (Hyla crucifer) were positively correlated with pH. This
latter species often breeds in stream inflows and outflows or along
the littoral zone of lakes, habitats also subjected to particularly
heavy acid loads as a result of snow melt (Clark and Euler 1981).
Bullfrog (Rana catesbeiana) and wood frog (Rana sylvatica) densities
were also reduced in acidic streams and ponds (Clark and Euler 1981).
Strijbosch (1979) reported a negative correlation between pH and
percentages of dead and moulded egg masses of frogs and toads in the
Netherlands.
Laboratory experiments have demonstrated that reductions in pH are
both directly and indirectly responsible for mortalities and
deformities found during amphibian embryonic development. Gosner and
Black (1957) studied the sensitivity of 11 species of frogs and toads
to conditions of depressed pH and found that the embryos were more
sensitive than adults. Frogs may undergo iono-regulatory failure due
to acidic conditions (Fromm 1981) similar to that reported for fish
(Leivestad and Muniz 1976; McWilliams and Potts 1978; Muniz and
Leivestad 1980; Packer and Dunson 1970). In the case of the cricket
frog (Acris gryllus) and northern spring peeper, an exposure of
embryos to water in the vicinity of pH 4.0 for a few hours resulted
in greater than 85% mortality. Beebee and Griffin (1977) noted
abnormalities in natterjack toad spawn exposed to low pH, and Noble
(1979) observed delayed development and embryonic mortality in the
leopard frog (Rana pipiens) at pH less than 4.75. The leopard frog
may be more sensitive to low pH than the latter study indicates.
Schlichter (1981) found decreased sperm motililty at pH values less
than 6.5 and the percentage of eggs which formed healthy embryos
decreased below pH values of 6.3. A similar study using the common
frog reported that sperm motility was reduced to 50% of maximum at pH
values of 6.4-6.7 and to 0% at pH values less than 6.0 (Gellhorn
1927, cited in Schlichter 1981).
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3-112
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Cook (1978) found no significant correlation between pond pH and
percent embryonic mortality in either the yellow-spotted salamander
or Jefferson's salamander studied in six ponds with mean pH values of I
5.3-5.6. In contrast Pough (1976) found heavy embryonic mortalities I
and deformities for both species in waters with pH values less than
6.0. Egg transplant studies suggest that yellow-spotted salamander •
eggs from acidic ponds are more tolerant to acidity than eggs from •
neutral ponds (Nielsen et al. 1977). While Hagstrom (1977) reported
the elimination of the common toad at pH values of 4.0-4.5, Cooke _
(cited in Beebee and Griffin 1977) found this species in waters of •
pH 4.2 and noted that tadpoles were able to tolerate this hydrogen •
ion level.
It is likely that other factors influenced by the acidity of the |
water may cause detrimental effects upon amphibian development. For
example, Huckabee et al. (1975) suggest that the combined effects of m
low pH and increased concentrations of aluminum, manganese and zinc •
may be the cause of the high mortality of shovel-nosed salamander
(Leurognathus marmoratus) larvae in Great Smoky Mountain National
P ark";•
Frogs, toads, and salamanders are important components of both
aquatic and terrestrial ecosystems. Orser and Shure (1972) reported •
that amphibians are among the top carnivores in temporary ponds and |
small streams, and are important predators of aquatic insects. In
turn, they serve as a high protein food source for other wildlife M
(Burton and Likens 1975b). Many birds and mammals depend heavily on •
these species for food (Burton and Likens 1975a; Cecil and Just
1979; DeBenedictis 1974).
3.7.7 Effects of Low pH on Fish
The purpose of this section is to review briefly how fish respond to |
low pH conditions. This will be done on the basis of documented
changes in fish population related to acidification, other field _
evidence and laboratory substantiation. For more comprehensive •
treatments of this subject, the reader is referred to reviews by ™
Fromm (1980), Haines (1981c) and Spry et al. (1981). In addition,
there is extensive literature available on laboratory studies (see •
Doudoroff and Katz 1950; EIFAC 1969), that were designed to elucidate •
mechanisms of pH toxicity. These laboratory results are reviewed, as
they are useful in explaining field observations and suggesting new •
directions for field studies. •
Results from laboratory experiments demonstrate how overall water ^
quality (i.e., hardness, ionic strength) can affect pH toxicity. For •
example, as ionic strength and water hardness increase, the short- ™
term sensitivity of fish to waters with pH values of 4 is decreased
(reviewed in Spry et al. 1981). The ameliorative effects of high B
Ca^+ and ionic strength appear most beneficial to early larval stages •
at intermediate pH values (^5). This is consistent with field
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observations that fish communities disappear from more dilute waters
at higher pH levels than they do from lakes with higher concentra-
tions of salts (Leivestad and Muniz 1976). In addition to hardness
and ionic strength, survival of fish in water of low pH is influenced
by the type of acid present (Packer and Dunson 1972; Swartz et al.
1978), temperature (Kwain 1975; Robinson et al. 1976), the level of
dissolved carbon dioxide in the water (Neville 1979), and by the
presence of metals (Baker and Schofield 1980; Swartz et al. 1978).
Salts are lost from plasma and body tissue of fishes exposed to low
pH conditions. Leivestad et al. (1976) found that Na+ and Cl~ in
blood plasma and K+ in muscle tissue declined in brown trout at low
pH levels. Increases in the concentration of Ca2+ enabled the trout
to regulate better ionic balances (Leivestad et al. 1980). Recent
studies by Saunders et al. (1982, in press) have shed light on
possible mechanisms affecting survival, growth, and the smelting
process in Atlantic salmon. Under low pH laboratory conditions it
was found that parr-smolt transformation was impaired, and ATPase
activity was lowered, resulting in a decreased salinity tolerance of
smolts. Salmon raised under low pH regimes (i.e., pH 4.2-4.7) were
found to have significantly lower plasma Na+ and Cl~ levels, which
was indicative of an impaired osmoregulatory ability in fresh water.
Field evidence suggests that the susceptibility to low pH appears to
be species-specific. From his studies of La Cloche Mountain lakes,
Beamish (1976) estimated the pH at which reproduction ceased in 11
species of fishes (Table 3-20). As well as interspecific differences
in sensitivity, variability in sensitivity has also been observed
among different strains of the same species (Robinson et al. 1976;
Swartz et al. 1978). However, it is likely that the acidification of
lakes and rivers in North America is proceeding too rapidly to enable
genetic selection for acidic tolerant strains to occur naturally
(Schofield 1976b).
Results of laboratory and field studies have demonstrated that some
species of fish are particularly sensitive to low pH levels in
certain reproductive stages (reviewed by Spry et al. 1981). Low pH
can inhibit gonadal development (Ruby et al. 1977, 1978), reduce egg
production (Craig and Baksi 1977; Mount 1973) affect egg and sperm
viability (EIFAC 1969; Menendez 1976) and inhibit spawning (Craig and
Baksi 1977; Menendez 1976). Embryonic development may also be
affected by low pH (Swartz et al. 1978; Trojnar 1977) and low
environmental pH can affect egg internal pH (Daye and Garside 1980).
Generally, fry appear less resistant to low pH than eggs (Spry
et al. 1981), and therefore fry may be particularly vulnerable to low
pH conditions associated with spring melt and storm events.
Hulsman and Powles (1981) conducted experiments on walleye eggs. The
eggs were incubated in situ in a series of small streams in the
La Cloche area of Ontario. The various sites ranged in pH from 4.60
to 6.72. Hatching success was significantly reduced in the clear
dilute streams with pH values less than 5.40.
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TABLE 3-20. APPROXIMATE pH AT WHICH FISH IN THE LACLOCHE MOUNTAIN
LAKES STOPPED REPRODUCTION (Beamish 1976)
6.0 to 5.5
5.5 to 5.2
5.2 to 4.7
4.7 to 4.5
Species
Smallmouth bass
Micropterus dolomieui
Walleye
Stizostedion vitreum
Burbot
Lota lota
Lake Trout
Salvelinus namaycush
Troutperch
Percopsis omiscomaycus
Brown bullhead
Ictalurus nebulosus
White sucker
Catostomus commersoni
Rock bass
Ambloplites rupestris
Lake herring
Coregonus artedii
Yellow perch
Perca flaveseens
Lake chub
Couesius plumbeus
Family
Centrarchidae
Percidae
Gadidae
Salmonidae
Percopsidae
Ictaluridae
Catostomidae
Centrarchidae
Salmonidae
Percidae
Cyprinidae
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3-115
One mechanism which appears to contribute to species extinction in
acidified systems is the failure of recruitment of year classes. In
a study of 38 La Cloche lakes, Ryan and Harvey (1980) reported
evidence of recruitment failure in yellow perch (Perca flavescens)
populations in the two lakes of lowest pH values: Patten Lake
(pH 4.1) and Terry Lake (pH 4.3). The age group composition of
yellow perch in Patten Lake is illustrated in Figure 3-40. Ryan and
Harvey (1981) also found evidence of reduced and missing year classes
of young fish in five populations of rock bass (Ambloplites
rupestris) in acid-stressed La Cloche lakes.
The absence of older individuals in populations of fish in some
acid-stressed lakes has also been reported (Harvey 1980; Ryan and
Harvey 1980). This effect is illustrated by the changes in age
composition of white suckers in George Lake, Ontario from 1967 to
1979 (Figure 3-41), Rosseland et al. (1980) also reported the
absence of post-spawning age perch and brown trout in three lakes
within the Tovdal River System, Norway.
In the field, there have been several reports of fish kills appar-
ently related to the low pH of rivers and lakes. In Scandinavia, for
example, Jensen and Snekvik (1972) reported mass mortality of
Atlantic salmon (Salmo salar), and Leivestad and Muniz (1976)
reported a brown trout (Salmo trutta) kill. Both fish kills have
been correlated with reduced water pH, although Al was not measured
in either case.
In North America, Harvey (1979) reported mortalities of several
species, primarily pumpkinseeds (Lepomis gibbosus) in Plastic Lake,
Ontario, during spring snowmelt runoff and pH depression. Surface
water pH was 5.5, while the pH of the major inlet stream was 3.8.
During the spring of 1981, some in situ bioassays were conducted in
Plastic Lake (Harvey 1981). Rainbow trout, Salmo gairdneri, were
placed in cages at four locations in Plastic Lake and at four
locations in the control, Beech Lake. Three nonmetal cages of
35 fish were situated at each location. No mortality occurred at any
of the cage sites in the control lake (pH 6.09-7.34, alkalinity
132-390 ;ieq/L). In Plastic Lake, however, mortality ranged from 12%
at the lake outlet site (pH 5.0-5.85) to 100% at the inlet site
(pH 4.03-4.09). Although aluminum concentrations were not measured
at the time of the 1979 fish kill and aluminum data for 1981 is not
yet available, total aluminum concentrations in Plastic Lake during
the 1979 and 1980 ice-free season varied between 9 and 30 ug/L in the
lake, and between 240 and 490 Pg/L in the major inlet.
3.7.8 Effects of Aluminum and Other Metals on Fish
Concentrations of metals can be elevated in acid-stressed lakes
(Beamish 1974a; Raines 1981c; Scheider et al. 1979b) because of
increased atmospheric deposition, increased mobilization from the
sediments and/or mobilization from the watershed (see Section 3.2.4).
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3-116
35
30
25
x:
0)
b 20
"o
0>
1 15
z
10
5
0
PATTEN LAKE
-
-
c$$w i i i
012345
Age in
•
1
1
1
1
1
678
Years
Figure 3-40. Age composition of yellow perch
(Perca
captured in Patten Lake, Ontario, pH 4
Harvey 1980).
1
Q
1
1
10
1
1
1
1
1
1
1
1
1
1
1
1
1
f lavescens)
.1 (Ryan and •
1
1
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3-117
w
il
E
3
•z.
0 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15
Age in Years
Figure 3-41. Changes in the age composition of the white sucker
(Catostomus commersoni) in George Lake, Ontario
(Harvey et al. 1981).
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3-118
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One of the most important consequences for fishes of watershed
acidification is the mobilization of aluminum from the watershed to
the aquatic environment (Cronan and Schofield 1979). Elevated levels
of aluminum in waters have been shown to have serious effects on fish
within the pH range normally considered not harmful to aquatic biota
(Baker and Schofield 1980). _
Spry et al. (1981) give a simplified description of the complex
chemistry of aqueous aluminum. The solubility of aluminum is minimal
at pH 5.6-6.0, increasing as pH increases or decreases outside this •
range (Figure 3-42). At pH greater than 5.5, soluble aluminum is •
mostly anionic; at pH less than 5.5 it exists increasingly as a
cation. The solubility of aluminum is apparently regulated by some •
form of aluminum trihydroxide solid, Al(OH)3(s), which is minimally ||
soluble at pH values of 5.6-6.0 (Driscoll 1980b). Fewer hydroxyl
ligands at lower pH allow the aluminum to become cationic, eventually •
becoming Al^+ at pH values less than 4.5 to 5.0. Cationic aluminum •
is able to form complexes with a number of ligands, including soluble
organics and fluoride, decreasing its toxicity (Figure 3-42) (Baker
and Schofield 1980; Driscoll et al. 1980). •
Laboratory studies have shown significant reductions in fish survival
at inorganic aluminum concentrations of 100 and 200 pg/L and greater
for white suckers (Catostomus commersoni) and brook trout (Salvelinus
fontinalis), respectively (Baker and Schofield 1982; Schofield and
Trojnar 1980). Inorganic aluminum levels as high as 600 yg/L have «
been measured in acidic Adirondack waters (Driscoll 1980b). Baker •
and Schofield (1980) note that fry exposed soon after initiation of
feeding and yolk sac absorption were more sensitive to elevated
aluminum concentrations than were eggs and sac fry prior to yolk •
absorption. They also found that the presence of aluminum actually •
mitigated the toxic effects of low pH to fish eggs. The survival of
brook trout and white sucker embryos through the eyed stage at pH •
levels below 5 was significantly better in treatments with aluminum •
than without. After hatching, brook trout fry were more susceptible
to aluminum at the extremes of the pH range tested (4.2 to 5.5) than
at intermediate pH levels (Figure 3-43). The greater susceptibility •
of fry at these extreme pH values may reflect a dual mechanism of ^
aluminum toxicity. At low pH, aluminum (probably Al-'"*") appears to
cause osmoregulatory stress and loss of salts from blood plasma H
(Baker 1981; Leivestad and Muniz 1981). At higher pH values (5.5), |
precipitation of Al(OH)3(s) damages the gills and leads to clogging
by mucous (Baker 1981; Schofield and Trojnar 1980). Baker and mt
Schofield (1980) also found that, at all stages, white suckers were •
substantially more sensitive to low pH levels and elevated aluminum
concentrations than brook trout.
Schofield and Trojnar (1980) suggested that levels of aluminum, •
rather than pH alone, may be the primary factor limiting survival of
brook trout stocked in Adirondack lakes. Muniz and Leivestad (1980) •
and Schofield and Trojnar (1980) suggested that mass mortalities of |
fish, observed during episodes of acidification in the spring, were
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3-119
100
80
60
CO
3
CO
40
20
10
14 15
Time (days)
Figure 3-42.
Percent survival of brook trout fry plotted as a
function of time in treatment waters at pH level 5.2
with no aluminum (control) or with 0.5 mg Al added per
liter with no additional complexing agents (Al) or with
0.5 mg fluoride/litre (Al + F) or with 30 mg (Baker and
Schofield 1980).
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3-120
+ 100
C
1
C
o
4-f
o
C
3
u_
CO
0}
(0
~G -100
3
CO
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o
'w
0)
0)
t_
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o
QC
•5 -200
0>
a
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CO
\
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X
4.0
Figure 3-43.
4.5
5.0
5.5
PH
Slope of the regression line of brook trout survival
(arcsin transformation) as a function of total aluminum
concentration at each pH level, plotted as a function
of pH level. A positive slope indicates presence of
aluminum improved survival: a negative slope indicates
detrimental effects of aluminum (Baker and Schofield
1980).
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3-121
most likely a result of elevated concentrations of inorganic
aluminum, mobilized from the soils by strong acids present in
snowmelt water. The former study demonstrated that pH declines alone
(to levels of pH 4.7-5.0) did not induce physiological stress in
fish, as determined from changes in plasma chloride levels. However,
associated increases in aluminum levels to 0.2 mg/L or more were
found to be sufficient to induce severe stress and eventual mortality
(Muniz and Leivestad 1980).
Aluminum levels in streams in the Adirondack Region of New York
State (Driscoll 1980b), in the Great Smoky Mountains National Park,
U.S.A. (Herrmann and Baron 1980), and in the Muskoka-Haliburton area
of Ontario (total aluminum levels from 1976 to mid 1978 ranged
between 5 and 1000 yg/L in 60 streams) (Ontario Ministry of Environ-
ment data from ongoing studies), fall within levels demonstrated to
be lethal to fish in laboratory conditions. However, as the
laboratory studies have demonstrated, the evaluation of aluminum as a
toxic element in acidified waters is not a simple function of total
concentration. In evaluating the survival of indigenous fish
populations one must consider the form of aluminum, the level of
hydrogen ion, the fish species present and their life history stage.
Other metals besides aluminum also occur at elevated levels in acidic
waters (Section 3.2.4). Harvey et al. (1982) reported increased
lakewater concentrations of manganese were associated with decreasing
pH for 50 lakes in the Wawa area of Ontario. They found Mn was
elevated when pH values were less than 5.0 and reached very high
concentrations in strongly acidified lakes. In the La Cloche
Mountain lakes, Mn was correlated inversely with pH and Mn declined
in acidified lakes in the Sudbury area following neutralization
(Harvey et al. 1982).
Manganese has been considered a relatively non-toxic element, and
thus toxicological data are very limited. Lewis (1976) determined
that manganese concentrations up to 770 yg/L had no effect on
survival of rainbow trout in soft waters with pH levels 6.9 to 7.6.
Concentrations of manganese in acidic waters have been measured up to
130 to 350 ug/L (Dickson 1975; Schofield 1976c). Available data
suggest that manganese levels, by themselves, have no apparent
adverse effects on fish, although Harvey et al. (1982) found elevated
Mn concentrations in the vertebrae of white sucker (Catostomus
commersoni) from acid lakes.
Although laboratory bioassays examining effects of zinc on fish are
numerous, none of these studies considered soft waters with pH levels
below 6. Chemical models predict that as the pH level declines, an
increasing proportion of the total zinc concentration should exist as
the free aquo ion (Stutnm and Morgan 1970). For many metals, the free
aquo ion (i.e., Me^"1") is considered the most toxic form (Spry
et al. 1981). This has not been confirmed to be true for zinc but
care should be taken in extrapolating bioassay data and maximum
acceptable toxicant concentrations (MATC) determined for pH levels
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3-122
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above 6 to conditions in acidic waters. For the most part, however,
lethal concentrations of zinc in bioassays are 10 times the zinc
concentrations found in acidic lakes (Spry et al. 1981). Sinley
et al. (1974) estimated that the MATC for rainbow trout (Salmo
gairdneri) exposed to zinc in soft, circumneutral water was between
140 and 260 yg/L. Benoit and Holcombe (1978), in life cycle •
experiments with fathead minnows (Pimephales promelas) in soft water, I
determined that the threshold level for significant adverse effects
on the most sensitive life history state was between 78 and 145 yg/L.
Zinc concentrations in acidic waters range up to 23 to 56 yg/L •
(Henriksen and Wright 1978; Norton et al. 1981a; Schofield 1978; Spry •
et al. 1981).
I
In some regions, concentrations of cadmium, copper, lead and nickel
(see Section 3.2.4) are also elevated in acidic lakes. Relation-
ships between pH levels and cadmium, copper, lead, and nickel •
concentrations, however, vary markedly between regions. High I
concentrations of these metals probably result primarily from
increased atmospheric loading and deposition, and occur principally
in surface waters in close proximity to pollutant sources (e.g., I
Sudbury, Ontario, Nriagu et al. 1982). Concentrations of some of I
these metals in lakes in the vicinity of Sudbury have been demon-
strated to have definite adverse impacts on fish and other aquatic •
biota (Conroy et al. 1976; Yan and Strus 1980). Excluding lakes |
within 50 km of Sudbury, acidic Ontario surface waters have concen-
trations of metals ranging up to about 0.6 yg/L Cd, 9 yg/L Cu, _
6 yg/L Pb and 48 yg/L Ni (Spry et al. 1981). Spry et al. (1981) •
reviewed bioassay data available and noted no significant adverse
effects on fish survival and reproduction at concentrations up to
0.7-11.0, 9.5-77, 13-253, and 380 yg/L for cadmium, copper, lead, and •
nickel, respectively. In general, concentrations of metals in acidic I
waters are below these "safe" concentrations (unless there is a local
source of metal emissions). However: (1) most of these bioassays •
were conducted in waters with pH levels above 6, and (2) the I
possibility for synergistic effects has not been evaluated.
In some regions, bioaccumulation of mercury in fish has been I
correlated with low pH levels in lakes. These elevated levels of ™
mercury in fish may have adverse effects on consumers (e.g., man or
fish-eating birds and mammals; Sections 3.7.12 and 5.2). However, no I
data have been reported to indicate that this bioaccumulation has any I
adverse effects on the fish themselves (Haines 1981c).
Survival of fish populations in acidic waters is determined primarily I
by levels of pH and inorganic aluminum (Baker 1982; Schofield and
Trojnar 1980). Although concentrations of a number of metals are _
increased in acidic lakes and streams, definite effects on fish have •
been demonstrated only for aluminum (except for lakes immediately ™
around Sudbury, Ontario). Other metals may play a lesser, but as yet
undefined, role. I
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3.7.9 Accumulation of Metals in Fish
3.7.9.1 Mercury
There is substantial evidence of the effect of pH on mercury content
in fish (Brouzes et al. 1977; Hakanson 1980; Landner and Larsson
1972). Bisogni and Lawrence (1973) and Jernelov et al. (1976) have
argued that one reason fish in waters of low pH contain more
methylmercury than fish in waters of comparable mercury contamin-
ation, but higher pH, seems to be that more acidic waters retain the
monomethyl-form of mercury in solution. It is, however, important to
recognize that pH is not the only variable which determines the
mercury burden in fish. Other factors include mercury availability,
level of bioproduction (i.e., lake trophic state), lake flushing
rates and lake/watershed drainage area ratio (Hakanson 1980).
Few data exist to link mercury concentrations in fish to lake
acidification. However, an increase in concentrations of mercury in
fish from 1970 to 1978 is evident in some lakes in the Adirondack
Mountains (Schofield pers. comm.). In Ontario, Suns et al. (1980)
sampled young-of-the-year and yearling fish for contaminant studies.
Their data (Figure 3-44) demonstrate increased mercury concentrations
with decreasing pH in lakes in the Muskoka-Haliburton area. At any
given pH level, however, the variation of mercury concentrations in
fish is substantial. For lakes with similar pH, the mercury
concentrations were higher in fish from lakes with a higher ratio of
drainage area/lake volume. This result implies that a quantity of
mercury from either direct atmospheric deposition or from watershed
leaching is influencing the concentrations in fish. Data for 1981
are shown in Table 3-21 (Suns 1982). In 1980, the survey was
extended to include adult smallmouth bass. Fish from six of the nine
lakes studied had average mercury concentrations above the Canadian
guideline (500 ng/g) for unlimited human consumption. In one lake
mercury concentrations in fish exceeded the U.S. guidelines of
1000 ng/g (Suns 1982).
Because of increased mobility and leaching under acidic conditions
and/or deposition, it is possible that metals other than mercury may
be accumulating in fishes. At present, however, the data base is
extremely limited (Haines 1981c).
In a survey of Ontario lakes by Suns (1982), yearling yellow perch
were analyzed for body burdens of lead, cadmium, aluminum, and
manganese. The data are shown in Table 3-21 and are summarized
below.
3.7.9.2 Lead
A significant (p less than 0.01; r = -0.74) correlation was found to
exist between lead residues in perch and lake pH. Mean lead residues
as high as 428 ng/g were found from Moot Lake (pH 5.5) and 403 ng/g
from Fawn Lake (pH 5.4).
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200
180
160
140
o>
c
120
CO
^
+*
0)
o
i 100
o
I 80
60
40
20
13
4.5
LAKE #, NAME
1. Duck Lake
2. Little Clear Lake
3. Harp Lake
4. Bigwind Lake
5. Nelson Lake
6. Chub Lake
7. Crosson Lake
5.0
5.5
6.0
PH
6.5
A = 0.63
p < 0.05
11
TWP. LAKE #, NAME
Minden 8. Dickie Lake
Sinclair 9. Leonard Lake
Sinclair 10. Heney Lake
Oakley 11. Cranberry Lake
Bowell 12. Healey Lake
Ridout 13. Clear Lake
Oakland 14. Fawn Lake
7.0
TWP.
McLean
Mo nek
McLean
Guilford
McCauley
Stanhope
McCauley
Figure 3-44.
Mercury concentrations in yearling yellow perch vs.
epilimnetic pH for selected lakes in Ontario (Suns
et al. 1980).
7.5
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3-126
3.7.9.3 Cadmium
3.7.10 Effects on Fisheries in Canada and the United States
3.7.10.1 Adirondack Region of New York
I
I
Although there is evidence that lead concentrations in water as low
as 8 ng/L can cause neurological disorders in fish (Davies et al.
1976; Hodson et al. 1978; Holcombe et al. 1976), no data are •
available to relate body-burden accumulations to any significant •
biological response.
I
A statistically significant (p <0.05; r = 0.60) correlation exists I
between cadmium residue levels and lake pH (Table 3-21). Little |
reference material is available at this time to evaluate the
environmental significance of these cadmium accumulations. However, *j
a laboratory study using relatively hard water (pH 7.5; alkalinity •
980 yeq/L) showed that 80 yg/L killed 50% of the test population of
young-of-the-year largemouth bass in 82 days. The same study _
discovered that 8 iig/L induced "abnormal behaviour" in the young fish I
in 12-week exposure (Clearley and Coleman 1974). The young bass •
average body-burden accumulations of cadmium were 38 ng/g after a
four month exposure to a concentration of 80 yg/L. Although it is fl
difficult to apply these laboratory data to field conditions, it is ||
apparent that cadmium residue accumulations in fish tissue from
Ontario lakes, particularly in the more acidic lakes, were consider- •
ably higher than accumulations observed under laboratory conditions •
to cause biological effects.
I
3.7.9.4 Aluminum and Manganese
No correlations between lake acidity and mean residue accumulations •
were apparent in the 1981 collections. It is likely that differences |
in lake complexing capacities influence aluminum availability for
uptake. Therefore factors other than pH and alkalinity will have to «|
be considered to evaluate fully residue accumulations. •
Moreau et al. (1982) compared the chemical content of opercula and
scales of brook trout from lakes in Laurentian Park classified by •
Richard (1982) as more acidic (Group 1, described in Section 3.7.10) •
with the same calcified tissue from brook trout from three nonacidic
lakes (Group 3, also described in Section 3.7.10). They reported ij
that the content of manganese, zinc and strontium was significantly ||
higher in the calcified tissue of brook trout from the acidic lakes.
1
I
The Adirondack region is one of the largest sensitive lake districts
in the eastern United States, and it is also receives the highest •
annual loading of wet sulphate. A recent inventory of Adirondack I
waters classified lakes by type of fishery supported (Pfeiffer and
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3-127
Festa 1980). These authors suggest that acidic deposition has
exerted the greatest negative impact on the brook trout fishery.
Brook trout are frequently the only game fish species present in the
many small headwater ponds located at high elevations in the
Adirondacks and particularly susceptible to acidic deposition.
It is difficult to evaluate exactly how many fish populations have
been lost from Adirondack waters as a result of acidification. The
Adirondack region encompasses approximately 2877 individual lakes and
ponds. Pfeiffer and Festa (1980) note that 180 Adirondack ponds that
formerly sustained brook trout populations (either naturally or by
stocking) no longer support such populations. It has not however
been formally demonstrated that all (or most of) these populations
extinctions occurred as a result of acidic deposition. For at least
a few lakes (reviewed in Pfeiffer and Festa [1980]) historic records
of fish population status, fish management procedures, and water
chemistry do suggest that population declines were associated with a
decrease in pH level and that alternative explanations for the loss
of fish other than surface water acidification seem unlikely.
Schofield (1976a) surveyed high elevation Adirondack lakes (total 214
lakes). For 40 of these lakes, historical data on fish and pH were
available (Figure 3-32). In the 1930s, only 8% of these lakes had pH
<5.0, 10% had no fish whereas in the 1970s, 48% had pH <5.0 and 52%
had no fish. In some cases, entire fish communities consisting of
brook trout, lake trout, white sucker, brown trout, and several
syprinid species apparently have been eliminated over the 40-year
period (Schofield 1976a, 1981, 1982).
The present-day distribution of fish in Adirondack lakes and streams
in relation to pH provides additional circumstantial evidence of the
impact of acidification on fish. For high elevation lakes, Schofield
(1976b, 1981, 1982) noted that the occurrence of fish was reduced at
pH levels below 5.0 (Table 3-22 and Figure 3-32). Brook trout occur
less frequently in lakes with pH <5.0, white suckers at pH <5.1,
creek chub at pH <5.0, lake chub at pH <4.5 to 5.0, and brown
bullhead at pH <4.7 to 5.0 (Schofield 1976b). About 50% of high
elevation lakes had pH levels below 5.0 in 1975 and 82% of these
acidic lakes were devoid of fish (Schofield 1976b). High elevation
lakes, however, constitute a particularly sensitive subset of
Adirondack lakes. It cannot be inferred that 50% of all Adirondack
lakes have pH<5.0, nor that all lakes currently without fish once
had fish and have lost their fish populations as a result of
acidification.
Indices of fish population status in Adirondack streams (sample of 42
streams) were also found to be positively correlated (p < 0.05) with
pH measurements (Colquhoun et al. 1980).
In addition to these observations of fish population status in
Adirondack waters as related to acidity, Schofield and Trojnar (1980)
examined the effect of water quality on fish stocking success. Poor
survival of brook trout fall fingerlings stocked into Adirondack
-------
TABLE 3-22. DISTRIBUTION AND FREQUENCY
DURING SURVEYS OF
BRACKETS
pH <4.5
Total lakes 16
% of total 7.1
No fish 16
% 17.2
Fish 0
%
Brook trout 0
f .80
Lake trout 0
%
f
Bullhead 0
%
f
White sucker 0
%
f .15
Creek chub 0
%
f
Golden shiner 0
% 15.0
f .15
Common shiner 0
f
Lake chub 0
%
f
Redbreast
sunfish 0
%
f
Common sunfish 0
%
f
REFER TO
4.5-4.99
95
44.2
74
79.6
20
20.0
16(26)
19.5
.72
0(5)
8(8)
16.0
.40
3(1)
8.3
.28
0(7)
3(4)
15.0
.12
9(2)
9.1
.05
KD
14.3
.05
0
0(1)
OF OCCURRENCE OF FISH
SPECIES
3-128
COLLECTED
1
1
ADIRONDACKS LAKES >610 METRES ELEVATION. NUMBERS IN
EXTINCT
5.0-5.49
36
16.7
2
2.1
25
25.0
18(1)
21.9
1.00
1(2)
7.7
0.4
11(1)
22.0
.44
7(1)
19.4
.73
5
18.5
.20
3
5.0
.09
0(1)
0
0
0
0
POPULATIONS
5.5-5.99
15
7.0
1
1.1
11
11.0
11
13.4
.77
4
30.8
.36
5
10.0
.45
8
22.2
.32
7
25.9
.64
1
40.0
.36
3(1)
27.3
.27
2
28.6
.18
0
1
16.7
.09
(Schofield
6.0-6.49 6
28
13.0
0
22
22.0
17
20.7
.89
2
15.4
.09
14
28.0
.64
7
19.4
.42
5(1)
18.5
.23
8
15.0
.16
1
9.1
.05
0
0
1
16.7
.05
1976b)
.5-6.99
22
10.2
0
19
19.0
17
20.7
1.00
4
30.8
.21
9
18.0
.47
8
22.2
1.00
8
29.6
.42
3
10.0
.67
3
27.3
.16
1
14.3
.05
3
100.0
.16
2
66.7
.11
>7.0
3
1.4
0
3
3.0
3
3.7
2
15.4
.67
3
6.0
1.00
3
8.3
2
7.4
.67
2
3
27.3
1.00
3
42.9
1.00
0
2
66.7
.67
TOTAL
215
93
100
82
13
50
36
27
20
11
7
3
6
i
1
•
1
1
V
i
i
••
i
i
i
i
i
i
1
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3-129
lakes was significant, (p <0.05) associated with low pH levels and
elevated aluminum concentrations.
Schofield (1982) summarized available data relating water acidity and
fish population status for the eastern United States. With the
exception of studies in the Adirondack region, very few of these
studies included comprehensive inventories of fish populations and no
adverse effects of acidic deposition on fish have been definitely
demonstrated. Discussions generally refer only to "potential
impact".
3.7.10.2 Ontario
More data on inland fisheries resource effects resulting from lake
acidification are available from Ontario than from any other province
in Canada. The case study of lakes in the La Cloche Mountain range
by Beamish and Harvey (1972) is best known. These lakes have a
naturally low buffering capacity and are only 65 km southwest of the
Sudbury smelters. Some of the lakes had no fish populations at the
time of the first survey, 1965-66; others had populations that were
endangered, and still others were apparently in a healthy condition
(Beamish 1976).
The fish community of Lumsden Lake (one of 68 examined) has been
studied for 14 years. The following chronology of fisheries losses
has been assembled by Harvey (1980) from his studies with Beamish
(Beamish and Harvey 1972), from provincial government fish capture
records dating to the early 1960s, and from observations by local
anglers and residents for some species prior to 1960:
1950s - 8 species present
1960 - last reported capture, yellow perch, Perca flavescens
and burbot, Lota lota
1960-65 - sport fishery fails (pH 6.8, Sept. 1961)
1967 - last capture of lake trout, Salvelinus namaycush and
slimy sculpin, Cottus cognatus
1968 - tagged population of white sucker, Catostomus
commersoni disappears
1969 - last capture of trout perch, Percopis omiscomaycus
and lake herring, Coregonus artedii
1971 - last capture of lake chub, Couesius plumbeus (pH 4.4,
Aug. 1971)
In their study, Beamish and Harvey (1972) also reported the loss of
fish from nearby Lumsden III, Lumsden II and O.S.A. lakes. They
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3-130
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1
interpreted these observations as evidence that the factor(s)
affecting the fishes of Lumsden Lake were probably widespread. They
also noted that both sport and nonsport fishes had disappeared from
the lakes, suggesting that overfishing was not responsible. The loss
of populations of lake trout, lake herring, white suckers and other
species was attributed to decreasing pH. Historical data available •
for Lumsden Lake indicated that in one decade (1961-1971) the lake pH I
had decreased from approximately 6.8 to 4.4. Measurements of pH from
1961 or earlier were available for eleven other La Cloche Mountain
Lakes, and corresponding 1971 measurements for these lakes indicated 4
that pH had decreased one to two units in the decade.
I
Beamish (1974a) also examined fish populations in O.S.A. and Muriel
Lakes. He found that few fish remained in O.S.A. Lake. While
several species were present in Muriel Lake, only the yellow perch
population appeared unstressed. A case history of another La Cloche «
Mountain lake, George Lake, was compiled by Beamish et al. (1975) for •
the years 1966 through 1973. They estimated that the pH of George *
Lake decreased at an annual rate of 0.13 pH units. Coincident with
the reduction in lake pH, populations of lake trout, walleye, burbot IB
and smallmouth bass were lost in this period. In 1973, most brown m
bullheads, rock bass, pumpkinseeds and northern pike did not spawn.
Mountains and the concomitant loss of fish populations. He also
examined other possible explanations for the response of fishes in —
these lakes. He concluded that decreased pH appeared to be the •
principal agent stressing the fish populations, as well as controll- ™
ing the concentrations of metals.
Examination of the age distribution of white suckers in George Lake I
in 1972 indicated no missing year classes and it was concluded that
no major reproductive failures had occurred prior to 1972 (Beamish et •
al. 1975). The pH of George Lake was measured colorimetrically In •
1960 as 6.5, ranged between 4.8 - 5.3 in 1972-73 and was 5.4 in 1979
(Harvey et al. 1981). In 1967, the white sucker population contained _
fish up to 14 years of age. By 1972, almost no fish were older than •
6 years. Sampling in 1979 revealed that 90% of the population was ™
composed of two- and three-year old fish (Figure 3-41).
•
Harvey (1980) also showed that the white sucker population of Crosson 0
Lake (pH 5.1; Muskoka-Haliburton) had a truncated age distribution
with few fish older than five years (Figure 3-45) compared with the m
age composition of white suckers in less acidic Red Chalk (pH 6.3) I
and Harp (pH 6.3) lakes. Such a comparison must be viewed with
caution due to the natural variability of age structure between
lakes. However a change to a similar age structure patten was •
observed, coincident with declining pH, in George Lake (Harvey et al. •
1981).
Kelso et al. (1982) have recently reported on a survey of 75 |
headwater lakes varying in size from 1.6 to 120 ha in the Algoma area
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3-131
W
il
«*-
o
i_
d>
E
3
Z
100
RED CHALK LAKE
0 1 2 3 4 5 6 7 8 9 10 11 12 13 14
Age in Years
Figure 3-45. Age composition of the white sucker population of
three lakes in the Muskoka-Haliburton Region of
Ontario (Harvey 1980).
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3-132
of central Ontario. Most were found to be poorly buffered with 65%
of the lakes having alkalinities less than 200 yeq/L, 26% less than
40 yeq/L and 8% less than or equal to 0 yeq/L. In 55 of the lakes
sulphate concentrations were found to exceed bicarbonate. None of
the eight lakes with alkalinity values less than zero were found to
contain any sport fish, including brook trout, the primary sport fish
in this area of the Province.
Minns (1981) analyzed the Aquatic Habitat Inventory data base of the
Ontario Ministry of Natural Resources (OMNR). This data base
contains conductivity, pH, lake morphometry and fish species presence
information for 6,393 Ontario lakes (as of September 1980, the time
of analysis). The lakes contained in the data base were assumed to
be representative of lakes in the area surveyed. Analysis of the
data base for the presence of dystropic lakes indicated that very few
were included and therefore their affect on the analysis would be
minimal. Using relationships beween alkalinity, conductivity and pH,
lakes were classified into categories in terms of their acidification
status and the results were extrapolated to areas represented by the
sample. Minns estimated that 1,200 lakes in the province are too
acidic to sustain fish communities (lake pH less than 4.7) and
approximately 3,500 other lakes are approaching that condition (lake
pH 4.7-5.3). Most of these lakes are situated in watersheds in the
region of Sudbury and are small (i.e., less than 10 hectares). Minns
suggested that esocid and most percid communities are not currently
at risk whereas the brook trout, lake trout and bass communities
represent the most vulnerable resources.
3.7.10.3 Quebec
Fisheries investigations in the province of Quebec have concentrated
in the Laurentian Park. To determine the relationship between the
level of acidity and fish productivity in these lakes, the Quebec
Ministry of the Environment sampled 158 lakes in the area. Water
samples were collected through the ice, three weeks after the
beginning of snowmelt in March 1981. Most of the lakes sampled were
headwater lakes ranging in size from 10 to 25 hectares, with brook
trout populations.
Richard (1982) classified the lakes into three groups using a
multivariate analysis. The variables accounting for the greatest
between group variance are described following:
Group
1
2
3
Number
of Lakes
23
65
57
pH
5.2
5.9
6.4
Alkalinity
(WS/L)
8.5
45.6
130.6
HC03-/SO
0.1
0.6
1.8
Total
,2- Aluminum
(yg/L)
230.0
143.8
71.2
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3-133
In each group of lakes the average annual yield, the angling effort
and the mean weight of the fish caught (from detailed daily records
prepared by all fishermen) were compared (Richard 1982). Only those
lakes with nine years of continuous exploitation were included in the
analysis (12 lakes in Group 1, 30 lakes in Group 2, 36 lakes in
Group 3). During the last four years of study (1978-81) the mean
yield from Group 1 lakes (the most acidic) was not statistically
different from that of Groups 2 and 3. This conclusion was corrobor-
ated by examination of data from 34 additional lakes that had been
fished continuously for from four to six years (Richard 1982).
Fisheries management practices within the Laurentian Park provide for
closure to fishing when angling success was reduced as defined by a
lower mean weight or lower number of fish caught, or when spawning
habitat was disrupted. Forty-four lakes were not included in the
analysis as they had been closed to fishing for one or more years
preceding 1981. The 44 lakes which were closed to fishing included
43.5% of the most acidic lakes (Group 1) as compared with 36.9% of
Group 2 lakes and 17.5% of the Group 3 lakes. This comparison
suggests lower productivity in lakes in Groups 1 and 2, the more
acidic and acid-stressed lakes, than in Group 3 lakes.
Although the frequency of fisheries management problems was higher in
the more acidic and acid-stressed lakes, one cannot assume a direct
cause-and-effeet relationship with low pH, but only a general
association between fish productivity, pH and the oligotrophic
conditions of these waters.
3.7.10.4 Nova Scotia
There are 37 rivers flowing through Nova Scotia for which there are
records to verify that they are (or once were) Atlantic salmon rivers
(Farmer et al. 1980). For 27 of these rivers, almost complete
angling catch records are available (annual reports from federal
fishery officers) from 1936. Of these 27 rivers, 5 have undergone
major salmon stock alterations since 1936 by dam construction/
removal, and/or extensive hatchery stocking. Watt et al. (1983)
examined the effect of low pH on angling by dividing the remaining 22
rivers into two groups, based on 1980 pH levels. For the 12 rivers
presently at pH values greater than 5.0, only one shows a statistic-
ally significant decline in angling success since 1936, another shows
a significant increase, and 10 show no significant trend. Of the 10
rivers with pH values less than 5.0, 9 show significant declines, and
one shows no significant trend.
To combine the data so as to form averages for the two groups, the
records were first normalized by expressing each river's angling
catch as a percentage of the average catch in that river during the
first five years of record (1936-40). These percentages were then
summed and averaged for each of the two pH groups. The results
(Figure 3-46) reveal virtually identical angling catches in the two
-------
200
O)
£100
CO
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O>
co
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CO
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•S 40
20
3-134
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•£ 10
8
Mean for 12 rivers with pH>5.0(1980)
Mean for 10 rivers with pH£ 5.0 (1980)
j_
_L
_L
JL
1935
1940
1945
1950
1955
1960
1965
1970
Year
Figure 3-46.
Atlantic salmon angling data normalized to facilitate
the comparison between high and low pH rivers. Each
river's catch was expressed as a percentage of the mean
catch in 1936-40 so as to give all rivers equal
weighting, and the two groups were then averaged by
year (Watt et al. 1982).
1975
19f
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3-135
groups until the early 1950s; after which the angling catches in
rivers of pH less than 5.0 declined, while the catch in rivers of pH
more than 5 continued to show no significant trend with time.
Factors other than pH (e.g., stream flows and sea survivals) also
affect the angling success. Variation from these other factors
should, however, affect both groups similarly. The apparent reason
for the difference in angling success between the two groups of
rivers is a difference in pH since the 1950s.
Historical water chemistry data are available for some of these
affected rivers from surveys performed in 1954 and 1955 (Thomas
1960). In the past 25 years, the pH of the Tusket River has
decreased from an annual range of 4.9-6.1 to 4.6-4.9; the Roseway
from a range of 4.4-6.4 to 4.3-4.5; the Jordan River from about 5.1
to a range of 4.4-4.6; the Medway River from a range of 5.5-6.5 to
5.1-5.8; and the Clyde River has decreased from 5.0 to 4.6.
Alkalinity values were below zero in the Tusket, Clyde, Roseway and
Jordan rivers in 1979-80 (Watt et al. 1983), but was greater than
zero during Thomas' study 25 years earlier. Although Thomas (1960)
sampled some of these rivers only once, his data on river pH suggest
that salmon reproduction in a few rivers may have been adversely
affected due to acidity by the early 1950s, consistent with the catch
data presented in Figure 3-47.
Within Nova Scotia, the pH of surface waters xs well correlated with
geology (Watt 1981). Seasonal variation in the pH of those rivers is
about 0.5 units, with the annual minimum occurring in mid-winter, and
a maximum in late summer. At present there are seven rivers with pH
less than 4.7 that previously had salmon but now have no salmon or
trout reproduction; 11 rivers are in the pH range 4.7-5.0, where some
salmon mortality may be occurring; and seven rivers are in the pH
range 5.1-5.4, which is considered borderline for Atlantic salmon
(Figure 3-48). Those rivers represent 2% of the total Canadian
habitat potential for Atlantic salmon, and 30% in Nova Scotia.
The numbers of salmon angled, recorded by Canadian federal fisheries
officers since 1936 in six Nova Scotia rivers are illustrated in
Figure 3-47. The Clyde River with a mean annual pH of 4.6 in 1980-81
has produced no angled salmon since 1969. Electroseining in the last
several years also produced no salmon. The Ingram River with a mean
annual pH of 5.0 (range 4.8-5.8) apparently still has a small
reproducing population; it was at one time a good producer of
Atlantic salmon. Federal fisheries officials consider this river to
be in imminent danger of losing its remaining stock. This river has
been identified by Canada Department of Fisheries and Oceans
personnel as a candidate for liming in order to create a refuge for
maintaining the gene pool of this stock.
One of the Nova Scotia rivers "threatened" by pH declines, the
Mersey, contains an Atlantic salmon hatchery. The Mersey watershed
has poorly developed soils, and its underlying geology is Devonian
granite. The mean total alkalinity of samples collected from the
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3-136
MIDDLE RIVER
Q
UJ
_l
O
z
i r ~~ii i
1935 40 45 50 55 60 65 70 75 80
Year
z
O
<
CO
300-
250-
200-
150-
100-
50-
TANGIER RIVER
DC
til
m
1935 40 45 50 55 60 65 70 75 80
Year
200-i
150-
10O-
50-
SALMON RIVER
1935 40 45 50 55 60 65 70 75 80
Year
INGRAM RIVER
"IIIII
193540 45 50 55 60 65 70 75 80
Year
EAST RIVER
40 45 50 55 60 65 70 75 80
1935
Year
CLYDE RIVER
193540 45 50 55 60 65 70
Year
75 80
Figure 3-47. Angling records for six Nova Scotia Atlantic coast
rivers with mean annual pHs (1980) <5.0 (Watt et al.
1983).
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3-137
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pH <4.7 (no natural salmon reproduction)
pH range 4.7 - 5.0 (some mortalities likely)
pH range 5.1 - 5.4 (fisheries threatened)
pH > 5.4 (no immediate acidification threat)
Figure 3-48.
The Altantic salmon rivers of the Maritimes have been
divided into 4 pH (estimated mean annual) categories
based on significance to salmon reproduction. Present
evidence indicates that salmon cannot reproduce at pHs
below 4.7. Juvenile mortalities of 30% or more are
expected in the pH range 4.0-4.7. Rivers in pH range
5.1-5.4 are considered threatened. Above pH 5.4 there
is no immediate acidification concern with regard to
Atlantic Salmon (Watt 1981).
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river in 1978-79 was less than 10 yeq/L, while mean pH was 5.2 (range •
of 4.9-5.4) (Farmer et al. 1980). In 1954-56 the river had a mean pH
of 5.8, with a range of 5.4-6.6, and a mean total alkalinity of —
48 yeq/L CaC03, with a range from 20 to 88 (Thomas 1960). Mean •
sulphate values have been estimated to have increased from 76 yeq/L
in 1954-55 to 158 yeq/L in 1978-79. During the period 1975-78,
mortality of Atlantic salmon parr reared at the Mersey hatchery I
typically occurred during the third and fourth weeks after first •
feeding. This higher-than-expected mortality was attributed to
increased acidity in spring river water supplying the hatchery. In •
1979, by treating the water with CaC03, the salmon fry mortality was •
reduced from 30% to 3% (Farmer et al. 1980). In 1980, the water was
again treated and produced the same increase in survival of parr. ^
Farmer et al. (1980) noted that, even though all rivers classified as ™
presently unsuitable for salmon historically sustained Atlantic
salmon populations, these rivers are all also naturally somewhat •
acidic and historically had relatively low fish production. Of the |
20 readings of apparent water colour (rel. units) (an indicator of
the presence of organic acids) presented for the 7 rivers classified m
"unsuitable" by Farmer et al. (1980), 16 were 100. For "threatened" •
rivers, only one of 21 readings was 100; the remaining readings
averaged 69. For rivers classified neither "unsuitable" nor
"threatened," and with pH readings above 5.5, the mean measure of •
apparent colour was 44. High degrees of colour are largely attribut- •
able to humates from peat deposits and bogs common in this area.
Inputs from these materials probably contribute to the low pH levels •
in "unsuitable" and "threatened" rivers. Historical records of pH in |
these rivers do, however, indicate that acidity has increased in
recent years. Watt et al. (1983) concluded "the Atlantic coast am
rivers of Nova Scotia have become more acidic over the past 27 years •
in response to increased acid loading in the precipitation." This
increase in acidity has been clearly correlated with declines in
populations of Atlantic salmon in the same rivers. •
3.7.10.5 Scandinavia
1
Hendrey and Wright (1976) reported that "acid precipitation has
devastated the salmonid fish in southern Norway." Massive fish kills tm
of adult salmon and trout have been reported in their river systems, •
usually occurring during the spring snowmelt or after heavy autumn
rains. An intensive survey of 50 lakes in southern Sweden showed
that inland freshwater species are also threatened. The decreases in •
pH have resulted in the elimination of Atlantic salmon from many •
Norwegian rivers in the past 20 years. Scandinavian scientists have
concluded that, directly or indirectly, the principal cause of the •
fish losses is acidification of the waters, due to acidic deposition. ||
Portions of Canada's Atlantic salmon fishery appear to have declined
as a result of acidification as has been experienced in Norway and .
Sweden. •
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3.7.11 Response to Artificial Acidification
While we know that the end-product of acidification includes the
disappearance of important fisheries, many of the early changes which
occur in acidified ecosystems are relatively unstudied. Furthermore,
it is not known whether declines in fish stocks are due singly or in
combination to the toxic effects of hydrogen ion, to hydrogen ion and
aluminum or other metals synergisms, to food-chain effects resulting
from elimination of critical species of animals and plants or
disruption of nutrient cycles.
A whole-lake acidification experiment was done in Lake 223 in the
Experimental Lakes Area, Ontario, in order to examine some of these
possibilities. The pH of the lake was progressively lowered from a
natural value of 6.5 to 6.9 (x = 6.7) to an average value of 5.1
by additions of sulphuric acid between 1976 and 1981. Detailed
monitoring of chemical, physical and biological changes, as well as
physiological and ecotoxicological studies, were done throughout this
period. Earlier biological results were summarized by Schindler
et al. (1980), Schindler (1980), Malley and Chang (1981), and
Schindler and Turner (1982).
Biological changes in the lake as it was artificially acidified and
the pH thresholds at which these changes occurred are summarized in
Table 3-23. The first changes which could have adversely affected
lake trout and white sucker populations occurred in 1978-79, when
populations of two species which are the usual prey of trout, fathead
minnow (Pimephales promelas) and oppossum shrimp (Mysis relicta),
collapsed. Despite these changes, no effects were detected in trout
populations. A succession of strong white sucker year-classes in
1978-80 and greatly increased abundance of pearl dace were adequate
food alternatives for trout. Apparently, the pearl dace partially
occupied the vacated fathead minnow niche, while the primary food
source of white suckers, benthic dipterans, increased in abundance
(Davies pers. comm.). In addition, the appearance of excessive
growths of Mougeotia in the littoral beaches probably provided
excellent nursery areas for sucker fry, but increased water transpar-
ency (Schindler 1980) perhaps made prey capture easier for trout.
Even though many changes have occurred in lower trophic levels,
juvenile and adult white sucker and lake trout populations have shown
little indication of stress, except for recruitment failures in the
very recent years of acidification, at pH values of 5.35 and below.
Up to 1981, populations of both species increased, and their growth
rates have remained high. Relative condition (a quantitative measure
of fish fatness) has decreased progressively for trout from 1977 to
1980, and for white suckers from 1978 to 1980, but this would be
expected due to the increased abundance of both species over the same
time period.
The relatively swift collapse of the fathead minnow population is due
to two factors. Firstly, a recruitment (year-class) failure occurred
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3-140
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3-141
in 1978 (pH -5.8). This agrees well with the results of Mount
(1973), who found that impaired reproduction of the same species
occurred at this pH in laboratory studies done at a variety of pH
values. Secondly, even under preacidification conditions, this
species had a very short life span of three years in Lake 223. Even
under natural conditions, during the second and third years of life
an extremely high natural mortality rate occurred, over 50% per year
(Mills pers. comm.), presumably caused in large part by trout
predation. Very few individuals remained after the second year of
life. Therefore, the failure of one year class in 1978 would leave
few spawning adults (age 2 and 3) the following year. Population
recovery was, therefore, almost impossible. The combination of
successive year class failures in 1978 and 1979 assured the rapid
disappearance of this species from Lake 223.
The thresholds observed for disappearance of key species and
appearance of others in Lake 223 agree well with observations made in
other acidified lakes.
For example, Mysis in Lake 223 disappeared in the same pH range as
benthic crustaceans with similar food habits disappeared in
Scandinavian lakes (0kland and 0kland 1980). Mougeotia epidemics in
Lake 223 began at almost the same pH values as in Swedish waters
(Hultberg pers. comm.). Recruitment failures in lake trout and
white sucker began in the same pH range that year classes began to be
absent in lakes near Sudbury and in Scandinavia (Harvey 1980; Muniz
and Leivestad 1980; Raines 1981b,c).
The Lake 223 results also demonstrate the danger of assessing
biological damage from acidification solely on the basis of game fish
populations. Major alterations to fish habitats and prey species
occurred several tenths of a pH unit above where initial damage to
lake trout was detectable, even with an extremely intensive study of
the trout population. The predation habits of lake trout appeared to
allow them to easily switch to pearl dace after the disappearance of
the fathead minnows which had been their normal prey.
In summary, the Lake 223 experiment clearly shows that alterations to
aquatic food chains begin at pH values slightly below 6.0. The
remarkable agreement between these whole lake experiments and
observational studies in Scandinavia and eastern North America
provides strong evidence that the observed declines in fisheries are
caused by acidification and not by other ecological stresses.
3.7.12 Effects of Acidic Deposition on Birds and Mammals
While birds and mammals are not affected directly by acidic depo-
sition they are vulnerable to changes in their habitat caused by
acidification, particularly to changes affecting the availability and
quality of their food. Although adults may continue to find
sufficient food in areas adjacent to their traditional nesting or
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breeding sites, they may be unable to obtain sufficient food to raise
young. In Scandinavia there have already been reports of such
effects on aquatic bird populations. Aimer et al. (1978) reported 4
that, "fish-eating birds, such as mergansers and loons, have been •
forced to migrate from several acidic lakes, with decreasing fish
stocks, to new lakes with ample food supply. In this way, many _
territories will become vacant and this will lead to decreasing •
stocks." While the extent of the problem has not yet been documented ™
in Sweden, Nilsson and Nilsson (1978) found a positive correlation
between pH and "water" bird species richness. "Water" birds were
defined as those species dependent upon open water, and included a
loon, and several species of waterfowl and gulls. From the results
of this study it was suggested that a reduction in young fish, a very «|
important food source for aquatic birds, may lead to low reproductive •
success and local extinction in some bird species (Nilsson and
Nilsson 1978). Eriksson et al. (1980) also proposed that reduced
reproduction of fish in acidified lakes may decrease the availability •
of fish of the size classes appropriate to young diving water birds. ™
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Losses of other aquatic organisms such as clams, snails, and
amphibians have been documented in acidified lakes and ponds (Section
3.7.6; Hagstrom 1977; Hall and Likens 1980; J. 0kland 1980;
K.A. 0kland 1980). While wildlife are largely opportunistic feeders, «
reductions of these organisms could affect the food availability for •
many wildlife groups such as waterfowl and semi- aquatic mammals.
The effects of changes in food and habitat will be difficult to
witness in the short term but, in time, breeding densities may •
decline and eventually productivity could fall in response to reduced 9
food availability.
The diet of the common loon (Gavia immer) is approximately 80 percent (
fish, the remainder being made up of crustaceans, molluscs, aquatic
insects, and leeches (Barr 1973). Because the food requirements of «
loons while rearing young are high and many of their food organisms •
are quite sensitive to acidification, the nesting densities of this *
species may be reduced. In eastern Canada, the common loon nests on
lakes throughout the susceptible terrain of the Precambrian Shield •
(Godfrey 1966). In central Ontario and Quebec as well as in the 0
Adirondack Mountains of the northeastern U.S., a number of lakes have
already been reported as devoid of fish as a result of acid loading •
(Beamish 1976; Schofield 1976a). Studies in New York indicate that |
loon productivity has remained high but nesting densities have
declined in the Adirondack region (Trivelpiece et al. 1979). To •
date, however, changes in loon populations in the Adirondacks have •
been interpreted only with respect to human disturbance; the probable *
role of food depletion has not been investigated. In Quebec,
fish-eating birds were found more often on the nonacid lakes •
(DesGranges and Houde 1981). The common merganser (Mergus merganser) |
and the kingfisher (Megaceryle alcyon) were observed only on those
lakes where the summer pH is higher than 5.6. In the vicinity of m
Schefferville, Quebec, important differences in numbers and composi- •
tion of lake-dwelling bird communities were found: a third as many
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3-143
species and a quarter of the total number of aquatic birds were
observed on lakes with pH less than 4.5 compared to lakes with pH
greater than 6.0 (DesGranges and Houde 1981). The situation is less
clear, however, for lakes of pH 4.5-5.5. It has been suggested that
the biomass of some forms of benthic invertebrates increases with low
to moderate inputs of acid because there are fewer fish predators
(Henrikson and Oscarson 1978; Eriksson 1979; Henrikson et al. 1980).
This may explain the larger number of invertebrate-feeding ducks
which are found on moderately acid lakes in southern Quebec
(DesGranges and Houde 1981) and in central Ontario (McNicol and Ross
1982).
Insectivorous birds such as swallows, flycatchers, and kingbirds may
be affected by lake acidity since this group of birds feed on
emerging insects and it is during the emergence that many insects are
most sensitive to high acid levels (Bell 1971). Because a number of
species of aquatic insects emerge in early spring during the peak of
acid input to lakes and ponds they are particularly vulnerable to the
effects of acid loading. It is also in early spring that the birds
have higher food requirements in nesting and raising young. In
southern Quebec, the tree swallow (Iridoprocne bicolor) was more
common during the breeding season in the vicinity of lakes of
pH >6.0 while in northern Quebec this species was not observed in
the area of lakes of pH <4.5 (DesGranges and Houde 1981). This was
also the finding from the studies of insectivorous birds in the
Killarney area of Ontario (Blancher 1982). The presence or absence
of these birds will largely be determined by the biota of the nearby
lakes.
Effects of acidification on lower life forms such as microorganisms,
essential to decomposition and nutrient cycling have been found
(Hendrey et al. 1976; Leivestad et al. 1976). A loss in productivity
at the base of the food chain due to decreased nutrient availability
could result in progressively larger reductions at each succeeding
trophic level. The implications for wildlife at the top of the chain
are a critical loss in biological production and severely reduced
carrying capacity of their habitat (Clark and Fischer 1981).
Increased solubility and mobility of metals from sediments have been
reported as a result of acidification (Schindler et al. 1980). The
higher concentrations of metals produced in lake waters have
important implications for biological organisms as described in
previous sections. Studies by Nyholm and Myhrberg (1977) and Nyholm
(1981) have implicated aluminum in the impaired breeding of four
species of passerines. Reductions in the reproductive success of
these birds was highly correlated with the distance of their nests
from acid-stressed lakes in Swedish Lapland. Breeding impairment was
manifested as abnormal egg formation producing thin and porous
shells. In addition, clutch size and hatching success of the
"affected" birds were reduced and egg weights were lower in the birds
closer to the acid-stressed lakes. The link between the acidified
lakes and the breeding impairment has been related to the high
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aluminum content of the limnic insects upon which the birds feed
(Nyholm 1981). Birds feeding closest to the stressed lakes have the
highest proportion of contaminated insects in their diets (Nyholm fl
pers. comm.; Eriksson et al. 1980). Similar findings of decreased |
egg size and weight were found for the eastern kingbird (Tyrannus
tyrannus) in the Killarney area of central Ontario (Blancher 1982). n
Although severe abnormalities in shell formation were not evident in •
the eggs examined in this preliminary study, egg porosity as measured
by the rate of water loss over the incubation period was negatively
correlated with pH. •
Elevated mercury levels have been found in fish in lakes with low pH
in central Ontario (Suns et al. 1980). In the Bohuslan area of •
Sweden, elevated levels of mercury were found in eggs of goldeneye |
(Bucephala clangula) (Eriksson et al. 1980b). Raccoons (Procyon
lotor) from the Muskoka area of Ontario support liver mercury levels M
of 4.5 ppm, a concentration five times greater than specimens from an •
area with nonacidified waters (Wren et al. 1980). Because neither of
these areas receives point source inputs of mercury, the sources are
believed to be leached from the watershed by acids or mobilized from •
sediments. Methylation of mercury has been related to the process of •
acidification and the formation of methyl mercury, a stable and
soluble form which readily bioaccumulates, is believed to be favoured
at low pH (Fagerstrom and Jernelov 1972).
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Results of a preliminary study of metal accumulation in the tissues ^
of moose (Alces alces) have established an age dependent increase •
in cadmium for tissues collected from 38 moose and 56 roe deer in ™
Sweden (Frank et al. 1981; Mattson et al. 1981). Aimer et al. (1978)
reported a 10-fold increase in levels of cadmium in acidified lakes I
on the Swedish west coast compared with those in nonacidified lakes •
in the same region. Cadmium may be accumulated in large concentra-
tions by some terrestrial and aquatic plants (Anderson and Nilsson M
1974; Hutchinson and Czyrska 1975), and therefore, metal contamina- •
tion of wildlife feeding on these plants may be an indirect effect of
acidic deposition. ^
A summary of potential effects on selected species of birds and ™
mammals dependent upon the aquatic ecosystem for their food and
habitat is presented in Table 3-24. This summary is based solely •
on feeding habits as research on the impacts of acidification on •
vegetation structure and productivity relating to wildlife habitat
is at a preliminary stage. •
3.8 CONCERNS FOR IRREVERSIBLE EFFECTS _
3.8.1 Loss of Genetically Unique Fish Stocks "
Loss of fish populations with specific gene characteristics from •
lakes and rivers may be an irreversible process. Over several |
thousand generations, most species appear to have evolved discrete
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3-146
3.8.2 Depletion of Acid Neutralizing Capacity
3.9 ATMOSPHERIC SULPHATE LOADINGS AND THEIR RELATIONSHIP TO
AQUATIC ECOSYSTEMS
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stocks adapted to similar, yet discrete and specific, habitats
(Loftus 1976). The basic unit of a stock is the gene pool, which —
is composed of a naturally sustained, genetically variable group of •
individuals, adapted through evolution to specific lake conditions. *
Surface water acidification is a stress that may reduce genetic
variability in populations of native fishes in sensitive areas. As •
an example, Beamish and Harvey (1972) documented the loss of gene |
pools of fish in acidified lakes in Ontario. The Ontario Ministry of
Natural Resources has attributed the extinction of lake trout •
(Salvelinus namaycush) in 27 lakes in the Sudbury-Temagami area to •
acidification (Olver pers. comm.).
A naturally evolved complex of stocks appears essential to utilize •
fully the productive capacity of waters. Therefore, it is important •
to recognize and preserve stocks (Haines 1981c; Loftus 1976; Ryman
and Stahl 1981).
I
Loss of discrete stocks may inhibit effective re-establishment of
naturally reproducing populations in waters undergoing rehabilitation m
and affect future opportunities for fisheries management. •
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Evidence seems to be conflicting as to whether the geochemical
alteration of watersheds due to acidic input should be viewed as •
irreversible, and, if so, on what scale. Irreversibility can be |
viewed most strictly as a failure to recover over geologic time; but,
for natural resource systems, an incomplete recovery to a prestressed _
or undamaged state over a few decades, for all practical purposes, •
may be regarded as irreversible.
Although irreversible reduction in acid neutralizing capacity of K
lakes and watersheds is one of the potential effects of acidic •
deposition, our present information base is insufficient to determine
its probability in impacted areas. •
3.8.3 Soil Cation and Nutrient Depletion _
The loss of soil cations, particularly Ca^+ and Mg2+, which can
lead to decreases in soil fertility (Overrein et al. 1980), is
another potentially irreversible consequence of watershed titration. •
However, the extent to which these cation losses represent a m
significant depletion of total available material is unknown.
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The previous sections have discussed chemical and biological changes •
observed in some surface water systems, including pH depression and
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3-147
associated effects over long-term, annual, seasonal and event-related
time series. Most of the results are consistent with the explanation
that they result from acidity associated with the 804 2~ and NC>3~
ions originating from atmospheric deposition. This section will
consider the significance of these levels of chemical alterations,
with a comparision of the annual deposition that could be associated
with acidification of the most sensitive streams and lakes. This
analysis requires consideration, not only of trends in surface water
and precipitation pH and sulphate concentration, but also of the
frequency and severity of brief periods during which much of the
response to the total acidic loading rate from runoff events is
expressed.
Emphasis has been placed on deriving as much information as possible
from comparisons of observed water quality and biological effects in
areas of varying deposition. These empirical observations integrate
many "unknowns" regarding soil water interactions which are impli-
citly taken into account by empirical comparisons. Loading rates
estimated from conceptual models of aquatic systems are compared to
the empirical observations. Such empirical approaches to support
environmental management are common. For example, flood structure
designs can be based on empirical relationships between discharge,
precipitation and physical characteristics of the watershed (Chow
1964). Vollenweider and Dillon (1974) used an empirical modeling
approach to set phosphorus loading criteria for eutrophication
control in lakes and reservoirs, and these have proven effective.
The following are the principal findings presented in previous
sections important in evaluating aquatic effects related to measured
acidic deposition:
1. Precipitation over most of eastern North America has hydrogen
ion concentrations up to 100 times those expected for distilled
water in equilibrium with atmospheric carbon dioxide.
2. Large quantities of sulphate and nitrate ions are deposited with
ff1" ions in precipitation in eastern North America.
3. Lakes in eastern North America with low alkalinities are
receiving elevated acid loadings. Such lakes, and their
associated streams, may suffer low pH and elevated metal
concentrations for short periods of time, particularly during
snowmelt and other periods of heavy runoff.
4. Stressed fish populations have been observed in lakes that
experience short-term low pH and elevated metal concentrations.
Mortalities of adult fish have been observed in one study lake
experiencing these conditions.
5. There are numerous examples of streams and lakes in Canada and
the United States that have experienced and are probably now
experiencing depletion of alkalinity. Fish populations that
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3-148
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survive short-term low pH conditions, will eventually be lost if
alkalinity is depleted and pH values fall below critical levels
causing successive reproductive failure. Long-term acidifica- m
tion has caused losses of fish populations in some lakes and •
streams.
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3.9.1 The Relative Significance of Sulphur and Nitrogen Deposition •
to Acidification of Surface Waters
Results presented in the previous sections have shown that four major |
ions of concern in acidic precipitation, (H+, NH^"1", N(>3~
and SO^- have some potential for altering lake and stream water g|
acidity. Soil and plant interactions with nitrate ions allow nitric •
acid to be largely assimilated by the terrestrial portion of the
watershed, except during periods of heavy runoff (Section 3.2.2)
(McLean 1981). In contrast, in many regions with poorly developed •
soils, that are limited in ability to neutralize acid, biological 9
uptake of sulphate is small in comparison to the mass balance of
sulphur (Harvey et al. 1981). Christophersen and Wright (1980)
reported that the sulphur export from a watershed in Norway was
essentially the same as the total input over the period November 1971
to October 1978. In a number of areas studied, where there exist no _
significant terrestrial sources or sinks of sulphur, SO^" is a •
conservative ion whose export to surface waters is directly related
to deposition in precipitation.
There are additional aspects to the issue of the dominant anion •
associated with the acidification of surface waters. These include:
1) the relative magnitude of S0^~ and NC>3~ in the rain and f
snow inputs, their variation during the year, and long-term
trends; M
2) the relative magnitude of the biological interactions of *
both anions in watersheds, as they are affected by
biological activity at different seasons and by changes in B
biomass over long periods; •
3) the production of alkalinity in terrestrial and aquatic •
systems when NC>3 is assimilated by plants; and •
4) the contact time of precipitation inputs with the water- ^
shed. •
Data presented in map form in Section 2 and other data presented by
Galloway et al. (1980g), McLean (1981) and by Harvey et al. (1981) •
indicate that acidic sulphur inputs exceed acidic nitrogen inputs V
over eastern North America on an annual basis. The net yield of
these anions to streams and lakes is predominantly SO^" on an annual •
basis (Harvey et al. 1981). Because nitrate reaches surface waters •
in small amounts relative to its loadings on an annual basis and does
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3-149
not accumulate in surface waters, its influence on long-term surface
water acidification is less than that of sulphate.
Further evidence that nitrate deposition is not principally respons-
ible for long term surface water acidification is given in
Table 3-25. Data for 21 headwater streams in the Muskoka-Haliburton
area of Ontario with a range of mean annual pH values from 4.08 to
6.18 show that as acidity increases, the relative importance of NOg
declines. The acid (H+) concentration exceeds the N(>3~ concentration
on a chemical equivalents basis for annual pH values of 5.5 or less,
so that lower pH values cannot be explained by the presence of nitric
acid. The E+/SO^~ ratios are also given for the same streams
(Table 3-25). At lower pH values, H+/S042" ratios increase.
The ratio is always less than one which indicates that the acid
concentration can be explained by the presence of sulphuric acid.
The SO^~/fiO^~ ratios range from 14 to 337 with a median value of
170, demonstrating the dominance of 8642- over NOg" in surface waters
in the Muskoka-Haliburton region (Jeffries et al. 1979; Scheider
et al. 1979c; and ongoing studies by Ontario Ministry of the
Environment).
Nitrate may be important on an episodic basis by adding to the pH
depression caused by sulphate. At Sagamore Lake, New York, nitrate
concentrations in the lake outflow increased during spring pH
depression, while sulphate concentrations did not increase (Galloway
et al. 1980g). Sulphate concentrations still exceeded nitrate
concentrations on an equivalent basis, even during spring runoff.
Uptake of nitrate ions by algae and aquatic plants results in the
production of alkalinity in surface waters (Goldman and Brewer 1980).
This has been shown to occur in one of the study lakes at Muskoka-
Haliburton. Reported increases in lake pH from 5.1 to 6.6 over the
summer were associated with decreases in nitrate concentrations by
photosynthetic processes, and this was given as the explanation for
the pH increases (Harvey et al. 1981).
The evidence available, and the published interpretations of that
evidence (Harvey et al. 1981; Overrein et al. 1980), lead to the
conclusion that, for surface water systems, increases in acidity are
the result of dilute solutions of strong acids reaching these waters.
Further, Harvey et al. (1981) following extensive analysis of
Canadian data and Overrein et al. (1980) following extensive research
in Scandinavia conclude that most of the acidity is due to the
changes observed in 804^" concentration attributable to sulphate and
sulphuric acid deposition (Harvey et al. 1981; Overrein et al. 1980).
Both sulphuric and nitric acid contribute acidity to surface waters
during periods associated with pH depressions and fish stress.
However, there is no strong evidence at present for anticipating any
appreciable reduction in long-term lake or stream acidification from
a reduction in nitrate inputs. In contrast, it is important to note
there is a strong correlation between between sulphate deposition and
surface water concentrations to suggest that a reduction in sulphate
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TABLE 3-25.
3-150
MEAN AND RANGE OF pH VALUES, MEAN H+/N03 , H+/S042
AND S042~/N03~ RATIOS (calculated as ueq/L) FOR 21
HEADWATER STREAMS IN MUSKOKA-HALIBURTON, ONTARIO 1976-
1980 [Data is from an ongoing study, methods and study
area as described in Jeffries et al. (1979) and Scheider
et al. (1979b)]
Stream
Dickie 11
Red Chalk 2
Dickie 5
Dickie 6
Dickie 10
Chub 2
Dickie 8
Harp 6A
Harp 5
Chub 1
Harp 3
Harp 6
Red Chalk 1
Red Chalk 3
Harp 3A
Red Chalk 4
Jerry 3
Jerry 4
Harp 4
Blue Chalk 1
Jerry 1
Mean
PH
4.08
4.30
4.34
4.35
4.59
4.82
5.03
5.19
5.34
5.41
5.64
5.77
5.81
5.95
5.95
5.96
5.98
6.07
6.08
6.16
6.18
Range
pH
3.53-5.61
3.68-4.81
3.71-4.76
3.74-5.05
3.92-5.10
4.12-6.08
4.04-5.87
4.34-6.39
4.66-6.60
4.48-6.61
4.89-6.39
5.20-6.90
5.19-6.69
5.17-6.65
5.30-7.30
5.28-6.71
5.27-6.67
5.49-6.55
5.29-6.90
5.71-6.62
5.58-6.74
H+/N03
(yeq/L)
93.60
60.00
58.30
60.20
25.90
23.90
12.60
9.57
2.49
5.49
1.05
0.83
1.74
0.21
0.29
0.24
0.51
0.46
0.15
0.67
0.04
H+/SO|~
(yeq/L)
0.457
0.188
0.318
0.297
0.119
0.071
0.049
0.028
0.017
0.019
0.009
0.007
0.009
0.006
0.004
0.006
0.004
0.003
0.003
0.003
0.003
SO|~/N03
(ueq/L)
245
265
233
247
170
236
284
337
145
232
156
130
174
34
100
37
134
118
57
198
14
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3-151
loading to watersheds would reduce the sulphate concentrations and
associated acidification of surface waters.
3.9.2 Data and Methods for Associating Deposition Rates with
Aquatic Effects*
The evidence available on the effects of acidic deposition on aquatic
resources indicates that present loadings rates are in excess of the
ability of watersheds to reduce the acidity for some lakes in some
areas. This section will explore the association between loading
levels of acids or sulphates and negative effects on the aquatic
environment. In the following analysis, it is implied that sulphate
deposition can be used as a surrogate for the acidifying potential of
precipitation.
The use of sulphate in precipitation as a surrogate for the acidi-
fying potential of deposition should not be interpreted to mean that
wet sulphate is the only substance potentially damaging to aquatic
systems. It is recognized that dry deposition of sulphate and SC>2,
and wet and dry nitrates contribute to the concentrations of acids.
Sulphate in precipitation is reliably measured and therefore, is used
here as a surrogate for the total sulphur deposition because dry
deposition cannot be measured accurately. Similarly, this surrogate
does not reflect the contribution of nitrate to acidity of precipi-
tation.
Surface water quality alterations fall into two categories:
1) short-term pH depressions during snowmelt or heavy rains,
and
2) long-term reductions in alkalinity, with corresponding low
pH values in surface waters throughout the year.
The length of time it takes for a lake to become acidic (alkalinity
reduced to zero or less) and the rate of change of water quality are
among the least well-defined aspects of the acidification phenomenon.
To date, the evidence available, based on sediment cores taken from
several areas (Section 6.3.4), suggests that acidification has
occurred and is occurring on the scale of decades.
* It is the view of the U.S. members of the Work Group that the
reliability of wet sulphate deposition values is uncertain and
therefore, any attempt to use them for analysis must be done with
great care. Examination of the data shows that: (1) limited
deposition data are available, and (2) annual variability in wet
sulphate deposition values can be large.
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Before the alkalinity of a lake or stream is totally depleted, it is
very likely that the system experiences short-term pH depressions
during periods of high runoff. Large temporal fluctuations in pH •
levels may represent a transition phase in the process of •
acidification.
The phenomenon of short-term pH declines is probably more common than |
long-term reductions in alkalinity (in terms of numbers of lakes and
rivers affected in North America). The chemistry of these events is ^
fairly well defined. The biological consequences of these events are •
known to be severe in some cases, but the relationship between
short-term pH depressions and effects on aquatic biota are not fully
understood. •
"In the second stage, the bicarbonate buffer is lost during longer
periods and severe pH fluctuations occur resulting in stress, •
reproductive inhibition and episodic mortalities in fish populations |
(transition lakes)" (Henriksen 1980). Damage to fish and other biota
as a result of short-term exposures to low pH and associated high _
metal concentrations has been demonstrated to occur in both •
laboratory and field studies (Section 3.7). Thus, summertime or ™
annual pH has questionable value for determining effects on organisms
of H+ or metals over a few days. The timing magnitude and duration I
of short-term increases in H+, associated with spring melt and V
storm events must, therefore, be included in an evaluation of
critical loading rate and episodic response relationships for streams •
and lakes. •
In summary, the short-term acute exposure or "shock" effects —
(including responses to aluminum) can take place in two to four days •
of exposure, with pH decreases in the order of 0.5-1.5 units; and •
these shock exposures can be expected to occur in waters with a broad
range of pH above the level at which chronic effects occur. B
The second category, long-term acidification, has altered a large
number of lakes in North America, but the percentage of lakes and a|
rivers with mean annual alkalinity of zero or less remains small. •
The biological responses to long-term acidification are, however,
more clearly defined and generally more severe than for short-term pH
declines. •
The acidity and chemical composition of aquatic environments are
affected by: (1) the acid neutralizing capacity of the basin; •
(2) the geologic and morphologic characteristics of the basin; and •,
(3) the acidity of the precipitation. Biological processes (e.g.,
production and decomposition) also have an effect on acidity. Models •
used to simulate the geochemical processes and aquatic ecosystem •
effects are not fully developed or validated at this time.
Development and application of detailed models will require detailed
information on basin geology, hydrology, and biotic interactions. •
These are unlikely to be available soon for widespread application. •
Therefore, at present, the relationships between acidic deposition
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3-153
and aquatic effects can be determined only in a general way. Some
data and phenomenological models exist that relate the behaviour of
lakes and streams to acid loading. These empirical observations and
models are discussed below.
3.9.2.1 Empirical Observations
Observed sulphate loadings and corresponding chemical and biological
observations for a series of study areas in North America and
Scandinavia are available.
The information in this section is drawn from a number of study areas
within eastern North America which are located on the Precambrian
Shield or on weathering resistant bedrocks. The surface water
studies have been initiated for several reasons, have started at
different times and are operated by different agencies. However,
each project contributes information relevant to the acidification
problem by comparison of results among and within the studies
themselves. In general, each project involves some highly detailed
work on a small number of watersheds and surface waters and less
detailed work on a larger study set. Within a given study area, the
surface waters and watersheds are usually chosen to cover as wide a
range of water quality and geology as is available.
The study area descriptions will give some appreciation for the
extent of the data base used in the empirical derivation of
loading versus chemical and biological effects.
SASKATCHEWAN SHIELD LAKES
More than 300 lakes in Northern Saskatchewan's Shield and Fringe
Shield regions have been sampled to assess the sensitivity of
lakes.
Deposition Annual Precipitation Annual Runoff
(kg S042-/ha.yr) (m) (m)
5a 1980 0.357 .100 - .200b
a CANSAP Measurement.
b Fisheries and Environment Canada 1978.
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Observed Characteristics
EXPERIMENTAL LAKES AREA, ONTARIO
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1. Alkalinity = -18.20 + 0.92 (Ca + Mg) (n=281,r=0.97)
Liaw (1982) indicating that the bicarbonate and Ca + Mg are
related by a 1:1 relationship and sulphate contributes very
little to the total ion balance. •
2. pH values range from 5.56 to 8.2, 39% <7.0 (Liaw 1982).
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The Experimental Lakes Area is situated in northwestern Ontario
on Precambrian shield granite. Approximately one-half the area
of Canada is Precambrian Shield. Within the study area there
are about 1,000 lakes, of which 46 lakes in 17 drainage basins •
have been set aside solely for experimental research. The •
inflows and outflow of Rawson Lake (a control lake) are
calibrated as well as 14 other watersheds. The project was
initiated in 1969 and is continuing a wide range of whole-lake •
chemical manipulations including the acidification of lakes with •
monitoring of chemical and biological parameters including fish
population studies. The results from this multi-faceted project •
are published in many scientific journals including two special •
issues of the Canadian Journal of Fisheries and Aquatic Sciences
devoted entirely to the Experimental Lakes Area (1971, ^
Volume 28, Number 2 and 1980, Volume 37, Number 3). I
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Deposition Annual Annual
(kg S042~/ha.yr) Fraction Time Precipitation Runoff
Period (m) (m) •
9.07a bulk 1972 0.69a 0.297a
10.8a bulk 1973 0.73a 0.354a g
5.9b wet 1980 0.51b 0.223a |
0.234a
0.15d •
Sum of Cations for 31 lakes 217 yeq/L _+ 25 (Standard
Deviation)
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a Schindler et al. 1976; see also Figure 3-16. •
b Barrie and Sirois 1982. |
c CANSAP measurement.
" Fisheries and Environment Canada 1978. •
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Observed Characteristics
1. No long term acidification or biological effects observed
in ten years of study (Schindler pers. comm; Can. J. Fish.
Aquat. Sci. 37(3); Can. J. Fish. Res. Board 28(2)).
2. Sulphate export from the watersheds is about equal to the
measured wet deposition (Schindler et al. 1976).
3. Lake alkalinity distributions for lakes in the Rainy River
district have fewer low alkalinity values than four other
Precambrian Shield areas in Ontario (Dillon 1982).
4. Lake pH values for a 109 lake survey ranged from 4.8 to 7.4
and averaged 6.5 (Beamish et al. 1976).
5. Lake sulphate concentrations ranged from about one-half to
about equal to the bicarbonate concentrations (Beamish
et al. 1976, Dillon 1982).
6. Filamentous algae are common in July and August but do not
dominate the algal population (Stockner and Armstrong 1971)
ALGOMA, ONTARIO
The Algoma region of Ontario is an area of 862,000 ha in
northcentral Ontario. From a chemical survey of about 85
lakes, Kelso et al. (1982) report results from 75 headwater,
nondystrophic lakes with watersheds undisturbed by recent
logging, fire or human settlement. Sampling was done in 1979-80
and included physical parameters, lake chemistry and
phytoplankton analyses on the entire lake set with benthic
invertebrate, sediment and fish tissue analyses done on subsets
of the 75 lakes.
The Turkey Lakes Project, situated within the Algoma region, is
an ongoing calibrated watershed study of 5 lakes and 20 water-
sheds, plus the outlet of the entire Turkey Lakes watershed
basin. Initiated in 1980, intensive chemical, hydrological and
biological studies are in progress including monitoring of
precipitation, air quality, forest effects, ground and soil
water, stream and lake chemistry.
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3-156
calculated from ion concen-
tration data from Kelso et al.
(1982) and precipitation data
from Barrie and Sirois (1982)
Barrie and Sirois 1982.
Fisheries and Environment Canada 1978.
Observed Characteristics
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Deposition Fraction Annual Annual •
(kg S042~/ha.yr) Time Precipitation Runoff
Period (m) (m) •
25 APN Turkey wet 1981 0.8a 0.50b
Lakes Station, _
(Barrie, pers. •
comm.)
28 1976
22 1977
32 1978
23 1979 _
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Sum of Cations for 75 lakes 285 yeq/L _+ 125 (Standard Deviation)
=™====™= I
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1. pH depression in streams during spring runoff up to 2.1 pH •
units with minimum values as low as 5.0 in streams with
summer alkalinities less than 400 peq/L (Keller and Gale
1982). •
2. Excess sulphate runoff is elevated about five times over
the remote areas of northwestern Ontario and Labrador •
(Thompson and Button 1982). Sulphate export from watersheds |
exceeds wet deposition indicating possible dry deposition of
sulphate. •
3. Of 75 headwater lakes surveyed, six had pH values of 5.3 or
less and the lowest value was 4.8 (Kelso et al. 1982).
4. Sulphate ions are the dominant anions (i.e., exceed •
bicarbonate) in lakes below pHs of about 6.5 (Kelso et al.
1982). •
5. In a survey of 31 headwater lakes (1.6-110 ha), the number
of lakes devoid of the 8 fish species reported in the area _
was observed to increase with decreasing alkalinity. The •
relationship between the presence of fish and pH in these ™
same lakes was weaker although a greater proportion of
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3-157
lakes of pH <5.5 were fishless than lakes of pH > 5.5
(Kelso et al. 1982). These observations are consistent
with the hypothesis that the biota in the surveyed lakes
have been adversely affected by changes in lake chemistry
but do not necessarily indicate causality (Kelso et al.
1982).
6. Aluminum and lead levels in 75 headwater lakes in Algoma
were elevated in lakes of lower alkalinity; mean total
aluminum levels of 53 yg/L was slightly greater than
aluminum levels in Muskoka-Haliburton waters (Scheider et
al. 1979a) and intermediate between concentrations found in
severely affected and slightly affected systems in Canada
and Norway (Kelso et al. 1982).
MUSKOKA-HALIBURTON ONTARIO
The study area in Muskoka and Haliburton counties of
southcentral Ontario encompasses an area of about 490,000
hectares within which are its 8 intensive study lakes and 32
calibrated watersheds, some of which have been calibrated since
1976. The watersheds vary in water quality and from low to high
pH. Twenty other lakes have been monitored on a seasonal basis
for a varying number of years. Many concurrent chemical and
biological studies are ongoing on the calibrated lakes as
summarized in Harvey et al. (1981). The results of these
studies have been reported in approximately thirty publications
in the primary scientific literature.
Studies of precipitation, deposition, air quality, soils,
groundwater, forests and precipitation throughfall are all being
carried out. A stream acidification experiment was started in
1982.
Studies of pH effects on fish and fish populations have been
intensified since 1979 by Harold Harvey of the University of
Toronto.
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Deposition Fraction
(kg S042-/ha.yr)
Sum
a
b
c
d
31 bulk
32 bulk
23 wet
29 wet
37 bulk
31 bulk
35 bulk
42 bulk
38 bulk
Annual
Precipitation
On)
0.8a
0.8a
1.2a
of Cations in Surface
Barrie and Sirois 1982.
Dillon et al. 1980.
Time Period
Aug76-Jul77
Aug77-Jul78
Aug76-Jul77
Aug77-Jul78
Jun76-May80
(Mean)
Jun76-May77
Jun77-May78
Jun78-May79
Jun79-May80
Annual
Runoff
(»>
0.45C
Waters 150-300
Fisheries and Environment Canada 1978.
Ontario Ministry of Environment, ongoing
3-158
Reference
Scheider et al.
1979a
Scheider et al.
1979a
Scheider et al.
1979a; Harvey
et al. 1981
Scheider et al.
1979a; Harvey
et al. 1981
Scheider & Dillon
1982
Unpublished4
Unpublished4
Unpublished4
Unpublished4
yeq/Lb
studies .
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3-159
Observed Characteristics
1. Severe pH depressions in streams and lakes with values as
low as 4.1 recorded (Jeffries et al. 1979).
2, Sulphate concentrations in lakes average about equal to the
bicarbonate concentrations (Dillon et al. 1980).
3. Manganese concentrations are elevated to about 50 yg/L
compared to about 3 yg/L at the ELA station (Dillon et al.
1980).
4. Aluminum concentrations (50 yg/L) are elevated over values
at ELA (Dillon et al. 1980).
5. Clear Lake, for which there are historical records, has
declined in alkalinity from 33 yeq/L in 1967 (Schindler and
Nighswander 1970) to between 2 and 15 yeq/L in 1977 (Dillon
et al. 1978), a reduction in alkalinity of greater than
50%.
6. Mercury concentrations are higher in fish from lakes with
low pH than from higher pH lakes (Suns 1982).
7. Unusually dense and extensive masses of filamentous algae
proliferate in the littoral zones of many lakes with pH
values of 4.5-5.5 (Stokes 1981).
8. Chrysochromulina breviturrita, an odour causing alga has
reached densities that have reduced the recreational use of
lakes for periods of time during the summer (Nicholls
et al. 1981). The species dominance appears to be a recent
phenomenon (within the past decade). This alga has been
shown to increase with decreasing pH in lake acidification
experiments (Schindler and Turner 1982).
9. Elemental composition of fish bones reported by Fraser and
Harvey (1982) showed the centrum calcium was reduced in
white suckers from lakes of pH 5.08 (King) and 5.36
(Crosson) compared to lakes of higher pH in the same area.
10. The white sucker population in Crosson Lake (pH 5.1) showed
a truncated age composition compared with the age
composition of the less acidic Red Chalk (pH 6.3) and Harp
(pH 6.3) lakes (Harvey 1980).
11. Adult pumpkinseeds (Lepomis gibbosus) and frogs have been
killed around the edges of Plastic Lake during spring melt
and acidification is the suspected cause. Inlet streams
had pH values as low as 3.85 (Harvey and Lee 1981).
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3-160
LAURENTIDE PARK, QUEBEC
Humid Alpine Lower Boreal Regions
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Elevated dome dominating the surrounding plateau. Elevation
varies from 500 to 1200 m asl with summit elevations of 1100 M
to 1200 m. It is comparable to the entire Laurentian plateau, •
although here there are very few lakes and the drainage pattern
is characterized by deep dissecting river valleys such as the
Jacques Cartier. •
The frost-free season is generally 80 days or less with a
growing season of about 140 days. Average annual rainfall, one •
of the most abundant in Quebec, ranges between 1200 and p
1600 mm.
On the upper slopes and summits, 85% of the surface is covered •
with glacial till of which two-thirds is less than 1 m deep,
while the other 15% consists of exposed bedrock (gneiss).
Low-lying areas are, for the most part, blanketed by sandy •
fluvio-glacial outwash deposits. A few organic deposits exist •
and are generally shallow, digotrophic and treed. Ferro-humic
podzols characterize the well-drained soils with little or no
ortstein to be found on excessively to well-drained sand soils.
I
The region, as defined by Thibault (1980), confirms early work _
completed by Jurdant and others (1968, 1972). The limits I
include all areas above 518 m. Jurdant (1968) and Lafond and *
Ladouceur (1968) characterized a distinct peripheral-band in the
central upland plateaus covered by balsam fir and black spruce •
moss forests and occasionally white birch stands. Forest •
regeneration after cutting or fire, is dominated by white birch
rather than trembling aspen. The central plateau supports a •
black spruce moss forest cover, but after cutting, regenerates •
and develops into a balsam fir Hylocomium, Oxalis forest (Lafond
1968). _
The more exposed summits in the region such as Mount Blie in the *
Malbaie watershed, support a scattered alpine cover dominated by
a heath, moss, and sedge complex and occasionally lichens. B
Humid Lower Boreal Region
This region, the Laurential foothills, is found between 47°30' •
and 50°00' N latitude and 67° and 75° W longitude. Mountainous
topography characterizes the region. _
Average growing season is about 150 days with a total annual ™
rainfall between 900 to 1000 mm. Due to altitudinal variations,
local climate conditions vary within the region. Lower •
altitudes, especially in the southern sectors are not as cold or |
as wet as conditions on the higher plateau, a difference of
200-300 degree days and an average rainfall 200-300 mm. •
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3-161
Near the foothills, crystalline Precambrian bedrock underlies
the region. Hillsides are generally covered by a thin (less
than 1 m) layer of till, with deeper deposits near the base and
scattered deposits on the upper slopes and summits.
Fluvio-glacial deposits characterize the valley floors of the
region. Ferro-humic and humo-ferric podzols are the dominant
soil formations.
Rowe (1972), Jurdant et al. (1972) and work completed using
provincial cover maps (MER-Ministe're d'Energie et des Resources)
were used to define the region. The limits as defined by
Thibault (1980) and Jurdant et al. (1972) regroup regions
considered by Jurdant as part of a large balsam fir-white birch
forest domaine. This domaine is characterized by a semi-dense
forest cover (60% crown closure nature, tree height greater than
21 m) of balsam fir and black spruce associated with white birch
and an absence of jack pine.
Rowe's forest region and the MER information confirmed the
region's limits. Mesic hillside conditions support balsam
fir-black spruce mass as well as black spruce-balsam fir mass
forest covers with white birch and white spruce associations.
Pure black spruce stands preferred either dry sites or poorly
drained hollows. White birch and to a lesser extent trembling
aspen associated with black spruce, balsam fir and white spruce
characterize the regeneration.
Except for a few isolated areas, the meridional sugar maple,
yellow birch, red maple, red pine, black ash and American elm
are not to be found in the region.
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3-162
Deposition Fraction Time Period
(kg S042-/ha.yr)
Reference
40
30/6 mo
10/6 mo
35
wet Apr79-Mar80
wet Apr79-0ct79
wet Nov79-Mar80
wet 1980
22.2
wet 28Sep81-27Sep82
Interpolated* from
Glass and Brydges
1982
Interpolated* from
Glass and Brydges
1982
Interpolated* from
Glass and Brydges
1982
Thompson and
Hutton 1982;
interpolated from
Barrie and Sirois
1982
Grimard 1982
Annual
Precipitation
(m)
1.14a
Annual
Runoff
(m)
0.95a
a Ferland and Gagnon 1974.
* Interpolations from existing deposition isopleth maps as a
basis for estimating deposition values can be in error.
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3-163
Major Cations in Peq/L
Ca
Mg
Na
K
Cond
Average
114.8
54.1
37.3
8.3
22.5
Standard Deviation
57.6
23.9
12.5
3.7
8.7
Observed Characteristics
1. The surface water pH is higher than the precipitation pH.
The pH of 152 lakes sampled in the last week of March 1981
and the first of April varied between 4.7 and 6.6 with an
average of 5.9 (Richard 1982).
2. The average content of sulphate in the lakes is of the
order of 80 yeq/L (Bobe"e et al. 1982; Richard 1982) and it
is higher or equal to bicarbonate.
3. The highest sulphate concentrations in lakes in Quebec and
the greatest alkalinity differences were observed in the
southwest. The lake water concentrations of sulphate and
the alkalinity deficits decrease to the north and east
(Bob€e et al. 1982).
4. There is a significant correlation (r = 0.76, p ^ 0.001)
between pH and total aluminum of the 152 lakes of Richard
(1982).
5. The Laurentide Park area is found in hydrographic regions
05 and 06 (Figure 3-13). Sulphate vs. £ [Ca] + [Mg] -
[alk] for these two hydrographic regions is found in Figure
3-14.
6. Compared to the pH of 1938-41, there is a greater
proportion of the lakes sampled 1979-80 in the classes of
pH 4.40-5.09, 5.10-5.79 and 6.50-7.19 amongst 5 pH classes
(Jones et al. 1980). Lakes in the two lowest pH classes
showed reductions in pH; the higher pH class increased
because of road salt and nutrient additions. The decline
in surface water pH tended to occur in the southern part of
the park.
7. In lakes continuously open to fishing for nine years prior
to 1982, average annual angling yield, angling effort, and
mean weight of fish caught in years 1978-81 were not
significantly related to lake pH. Management policies
within the Park provide for closure of a lake to fishing
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3-164
and 1.2 times higher in the population of the three more
acidic group of lakes comparatively to the three non-acidic
group of lakes (Moreau et al. 1982).
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when angling success is reduced below projected levels.
The 44 lakes which were closed to fishing over the nine
year period included 43.5% of the most acidic lakes •
(Group 1, mean pH 5.2); as compared with 36.9% of Group 2 •
lakes (mean pH 5.9) and 17.5% of the Group 3 lakes (mean
pH 6.4). Although a direct cause-and-effect relationship
between fish productivity and pH has not been established,
the greater number of closures in the more acidic lakes
suggests a lower productivity in these waters (Richard •
1982). •
8. The concentrations of manganese, zinc and strontium in the
opercula of Salvelinus fontinalis are respectively 1.6, 1.3 •
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NOVA SCOTIA
The Nova Scotian River Study by Watt et al. (1983) encompassed
the approximately 500 km long Atlantic coast of Nova Scotia
which is underlain by granite on about one-half of the mainland. B
This study of 23 rivers which historically supported salmon •
fisheries reports results of monthly monitorings from June 1980
to May 1981, with certain rivers studied as long as 10 years. •
An historical comparison of five of these rivers with data •
collected in 1954-55 (Thomas 1960) pH, alkalinity, and major ion
concentration data was made. Fisheries data for the past 45 _
years was available for 22 of the rivers and Watt et al. (1983) •
related angling success to current water chemistry and ™
geological factors. Within Kejimkujik National Park, central
Nova Scotia, an ongoing study involves three calibrated lakes. I
Kerekes (1980) reported results for these lakes for the •
June 1978 - May 1979 period. From this study a chemical budget
is available for the Mersey River (the outflow of Kejimkujik •
Lake), which is included in the fisheries data set of Watt •
et al. (1983).
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3-165
Deposition Fraction Time Period
(kg S042-/ha.yr)
Location
44
22
19
22-29
22
17
27
32
31
18.12
13.18
29.01
21.27
22.50
total
wet
wet excess
Jun78-May79
wet and 1977-79
dry excess
1981
1980
1980
1979
1978
Feb78-Dec80
Nov77-Dec80
Oct77-Nov79
May78-Dec80
Oct77-Mar80
Kejimkujik,
Kerekes (1980)
Interpolated*
from Figure 3,
Underwood (1981)
Kejimkujik3
Kejimkujikb
Truroc
Truroc
Truroc
East River
St. Marysd
Cobequidd
Bridgetown^
New Rossd
Kemptvilled
Annual Annual
Precipitation Runoff
(m) (m)
1.2e 1978 1 mf
1.6e 1979
1.2e 1980
1.40 June 1978 - May 1979
1.46^ long-term average
Sum of Cations for 41 lakes and rivers 59 _+ 17 ueq/L
(Standard Deviation)
a
b
c
d
e
f
Barrie pers. comm.
Barrie et al. 1982.
Truro CANSAP received a fair rating in the siting
assessment (Vet and Reid 1982) and the station is being
moved (Barrie pers. comm.).
Underwood 1981 and Underwood pers. comm.
Barrie and Sirois 1982.
Fisheries and Environment Canada 1978.
Interpolations from existing deposition isopleth maps as a
basis for estimating deposition values can be in error.
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3-166
Observed Characteristics
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1. Precipitation pH is generally lower than the pH of the •
runoff water. High runoff is associated with the lowest pH •
values in river waters. The lowest mean monthly values in
rivers generally occur in winter (Watt et al. 1983). •
2. Sulphate is the dominant anion in three study lakes of
pH 5.4, 4.8 and 4.5 (Kerekes 1980) and was highest in the _
two coloured lakes with lowest pH. •
3. Excess sulphate export from the watersheds are elevated
above those of remote areas by a factor of about 4 •
(Thompson and Button 1982) and sulphate export exceeds the •
measured wet deposition indicating possible dry
deposition. •
4. pH data are available for four rivers (corrected for flow)
and 1980-81 values are less than 1954-55 by 0.24 to 0.79 _
units. The current bicarbonate concentrations are lower •
and sulphate and aluminum concentrations are higher than ™
historical values (Watt et al. 1983).
5. Two rivers (St. Mary's and Medway) had the lowest pH values •
and highest excess sulphate loads in 1973. Similar changes
in pH and excess sulphate were noted for two Newfoundland •
rivers (see Figure 3-30). |
6. Long-term (five years or greater) records for pH, calcium —
and sulphate from eleven rivers in Atlantic Canada were •
fitted by time series models. Five of eight sensitive *
rivers decreased in pH and the other three did not change,
while none of four insensitive rivers decreased. B
Relationships between trends in pH and Calcium and sulphate 8
indicate that, conceptual models applied satisfactorily for
pH and in only a limited number of cases for calcium and
sulphate (Clair and Whitfield 1983).
7. Salmon catch data for 22 rivers which have not been _
affected by watershed changes or salmon stocking, have been •
recorded from 1937 through 1980. As a group (n = 10), •
rivers in the pH range 4.6 - 5.0 have reduced salmon stocks
as reflected by a significant decline in angling catches
over this time. Collectively, rivers with current pH
values >5.0 do not show any significant trend in salmon
catch over the past 45 years (Watt et al. 1983). The •
absence or reduced abundance of Atlantic salmon in 17 •
rivers was corroborated by electrofishing surveys in
1980-82 (Watt et al. 1983).
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3-167
8. Diatom assemblages in four Halifax study lakes shifted
toward more acid tolerant species between 1971 and 1980
(Vaughan et al. 1982).
BOUNDARY WATERS CANOE AREA AND VOYAGEURS NATIONAL PARK,
MINNESOTA
The Boundary Waters Canoe Area Wilderness (BWCA), a wilderness
unit within the Superior National Forest (Minnesota) and located
along 176 km of the Minnesota-Ontario border. The area varies
from 16 to 48 km in width. Over 1,900 km of streams, portages,
and foot trails connect the hundreds of pristine, island-studded
lakes that make up approximately one-third of the total area.
Most of the BWCA is included within the Rainy Lake basin, except
for the eastern section, which is part of the Lake Superior
watershed. Of a park total of 88,800 ha, several thousand of
the 34,700 ha of recreational water in the VNP were created by
dams, leaving 54,080 ha of land. The park has 31 named lakes
and 422 unnamed swampy ponds larger than 2 ha. The BWCA has a
surface area of 439,093 ha patterned by 1,493 lakes greater than
2 ha, and over 480 km of major fishing and boating rivers in
addition to numerous streams and creeks (Glass and Loucks
1980).
Filson Creek watershed is approximately 13 km southeast of Ely,
Minnesota. Filson Creek drains 25.2 km^ and flows north and
west to the Kawishiwi River. Included in the watershed are
Omaday and Bogberry Lakes and one tributary, designated South
Filson Creek for this study. South Filson has a 6.3 km^
drainage area and no significant lakes.
About 60% of Filson Creek watershed is covered by mixed upland
forest, 30% by wetlands and lakes, and the remainder by planted
or natural stands of pine. Wetlands surround the lakes.
The precambrian bedrock is mostly troctolite (a pyroxene-poor,
calcic gabbro) and other igneous rocks of the Duluth Complex.
The northern 10% of the watershed is underlain by the Giants
Range granite. A mineralized zone along the contact between the
granite and the Duluth Complex contains copper and nickel
sulfide minerals. The watershed has no carbonate rocks.
Bedrock is at the land surface in about 10% of the watershed.
Most of the watershed is covered by drift generally less than
1 m thick. Its mineral composition reflects the underlying
bedrock types. The total thickness of drift and peat under the
wetlands can exceed 15 m. The peat in most of the wetlands is
fibric, herbaceous, and partly decomposed (sapric) below about
0.75 m (Seigel 1981).
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3-168
Deposition Fraction
(kg S042-/ha.yr)
10-15 wet
13 wet
1 . 6 snow
17 bulk
17.2 wet
Time Period
1976-78
1981
1978
(snow season)
Nov76-0ct77
1980
Reference
Glass and Loucks
1980
NADP 1981-83
(Marcel site)
Glass 1980
Siegel 1981
NADP 1981-83
16.6
14.8
wet Apr78-May79
Wet Apr78-May79
(Marcell site)
Total NE Minn.,
Eisenreich et al,
1978
Heiskary et al.
1982 (Hovland
site)
Observed Characteristics
1. No known chemical or biological effects in lakes (Glass
1980; Glass and Loucks 1980).
2. Most of BWCA lakes surveyed have pH values <6.0 and 36.5%
had CSI >3 (Glass 1980; Glass and Loucks 1980).
3. Of the 290 sites sampled 50.5% had alkalinity values
between 40-199 ueq/L no lakes had alkalinity values less
than 40 yeq/L (Glass 1982; Glass and Loucks 1980).
4. Filson Creek watershed retained 10.6 kg S042~/ha.yr
of 17 kg S042~/ha.yr bulk (Siegel 1981)
5. S042~ increased from 2 to 14 mg/L and [H+] from pH
values of 6.6 to 5.5 during snowmelt (Siegel 1981).
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NORTHERN WISCONSIN
Northern Wisconsin is a region in which a collapsing glacial
mass left deep outwash sands and coarse tills interspersed with
ice-blocks. The study area encompasses portions of seven
counties in the Upper Wisconsin River Basin. Water covers 17%
of the area. The area has had a 30% increase in population over
the last decade, much of which has occurred along lakeshores.
Although only 3% of the total land area is developed,
approximately 40% of the lake shoreline is in residential land
use.
About 90% of the land surface in the region is now forested. A
century ago the upland vegetation was dominated by white pine,
hardwoods and hemlock, but most of it was removed during logging
and subsequent burning in the late 1800s and early 1900s.
Regrowth of aspen, birch, mixed hardwoods and a few conifers has
taken place now, much of it since 1920. Black spruce is common
on the wet, peat areas. The sands and sandy loams in the
surface layers have produced mostly acid soils (commonly pH
4-5), with low cation exchange capacities (10 meq/100 g) and low
base saturation (10-30%). The upland soils are primarily sands
and sand loams with peatland soils in the depressions. Total
concentrations of calcium and magnesium in these soils are
typically 1-2 meq/100 g.
The igneous and metamorphic bedrock underlying these northern
Wisconsin counties is part of a southern extension of the
Precambrian Canadian Shield. The principal bedrock type is
granite, with lesser amount of diorite, schist, gneiss,
quartzite, slate and greenstone. The bedrock is overlain by the
glacial drift, the most recent of which was deposited during the
Wisconsin glaciation. Drift thickness ranges between 10 and
70 m with an average slightly greater than 30 m. The drift is
low in calcareous material, calcareous pebbles are found only in
the deeper, older drift. Essentially all groundwater
contributions to lakes and streams follows a path through the
glacial drift. Because most of the lakes occur in pitted
glacial outwash or end moraines, they are generally shallow,
averaging about 10 m in maximum depth and rarely exceeding 30 m.
Consequently, virtually all of the lakes in this study area are
situated well above bedrock, encased in thick glacial deposits.
The recent pH of the rainfall has averaged 4.6 annually compared
with an estimated 5.6 in the middle 1950s. The climate is cool
and wet, with mean July temperatures of 19°C and January
temperatures of -11°C. The lakes commonly are ice-covered from
late November to late April (Schnoor et al. 1982).
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3-170
(Trout Lake)
17 wet (71 cm) 1980 NADP 1981-83
(Trout Lake)
16 wet (84 cm) 1981 NADP 1981-83
22 bulk 1981 Becker et al.
1982
Precipitation Runoff
(m) (m)
.80 .30
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(kg S042-/ha.yr) |
17 wet (68 cm) 1981 NADP 1981-83 I
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Annual Annual •
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Observed Characteristics
1. Median alkalinity for 117 seepage lakes sampled was 39
yeq/L. Conductivity and colour for the same lakes was 21 I
yS/cm and 8 Pt units (Eilers et al. 1982). For 409 total •
sites, 25.4% had alkalinities <40 yeq/L and 22.7% had
alkalinities between 40 and 199 yeq/L (Glass 1982). •
2. Two separate comparisons of present chemistry with the 500
Wisconsin lake survey of Birge and Juday (1925-41) have _
found that most lakes have significantly higher pH, I
alkalinity and conductivity (Bowser et al. 1982; Schnoor et "
al. 1982). Approximately 20% of lakes sampled had pH
declines but the differences were not statistically •
significant. •
3. Hydrologic type appears to control alkalinity. Median •
values of pH (6.4) alkalinity (39 yeq/L) and conductivity •
(21 ymohs) were found in seepage lakes (no defined inlet or
outlet) (Eilers et al. 1982; Schnoor et al. 1982). _
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3-171
ADIRONDACK MOUNTAINS OF NEW YORK
"As a result of extensive glacial activity, the Adirondack
region of northeastern New York State contains a vast and varied
ponded water resource. The most recent count adapted from a
1979 inventory of the Adirondack ecological zone (Pfeiffer 1979)
reveals that there are approximately 2,877 individual lakes and
ponds, encompassing some 282,154 surface acres. The New York
State portion of Lake Champlain, 97,000 acres, is purposely
excluded from this summary since its low elevation waters are
not considered to be representative of the Adirondack uplands.
Average size of ponded waters included in this inventory
approaches 98 acres and ranges from those of less than one acre
to 28,000 acre Lake George." (Pfeiffer and Festa 1980)
The Integrated Lake-Watershed Acidification Study (ILWAS)
selected three forested watershed areas (Panther, Woods and
Sagamore) in the Adirondack Park region of New York for field
investigation. The watershed areas contain terrestrial and
aquatic ecosystems having physical, chemical and biological
characteristics which distinguish one area from another. Lake
outlets are the hydrologic terminal points of all three
watersheds. The study watersheds are within 30 km of each
other. Runoff in Panther and Woods watersheds drains directly
to the lakes without extensive steam development. Sagamore Lake
receives the majority of its inflow through a drainage system of
bogs and streams. All watersheds contain mixtures of coniferous
and deciduous vegetation.
Panther Lake sits on thick till rather than bedrock. The
stratigraphy of the till is typically, from top to bottom, sand,
sandy till, silty till, and clay till overlying bedrock. The
till in Woods Lake basin is primarily sandy till with an average
depth of three metres. Panther Lake basin has two till units, a
sandy unit and a clay-rich unit; the two units together may be
60 m deep in places. Sagamore Lake basin has four units - a
loose sandy unit, a more compact sandy unit, a silt-rich unit,
and a clay-rich till. A thick sand deposit greater than 30 m
deep, at the site of a glacial meltwater channel, is present
near the inlet to Sagamore Lake.
High runoff periods typically occur during snowmelt. A winter
thaw has been observed in January and February. A larger spring
melt occurs in March and April. During the summer and fall,
occasional storms may also generate high runoff.
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Deposition
(kg S042-/ha.yr)
26.4
29
34-37
39-43
40.03
5.38
39.40
6.19
32.92
5.71
Fraction
wet
wet
bulk
bulk
wet
dry
wet
dry
wet
dry
Time Annual
Period Precipitation
(m)
1981 1.02
1980
1965-78
1965-78
Jun78-May79 1.25
Jun78-May79
Jun78-May79 1.21
Jun78-May79
Jun78-May79 .98
Jun78-May79
3-172
Reference
NADP 1981-83
(Huntington site)
NADP 1981-83
(Huntington site)
Peters et al.
1981 (Canton
site)
(Hinckley site)
Johannes et al .
1981
(Wood ' s Lake -
ILWAS)
Johannes et al .
1981
(Panther Lake -
ILWAS)
Johannes et al .
1981
(Sagamore Lake -
ILWAS)
1
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•
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1
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•
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3-173
Summary of 13 Years (1965-1978) Precipitation Data (Mean + S.D.)
(Peters et al. 1981)
Precipitation
Site (cm/yr) SO^2" (ueq/L) N(>3~ (yeq/L)
Canton 94 _+ 8 0.104 Hh 0.057 0.033 + 0.034
Hinckley 129 + 52 0.084 + 0.039 0.027 + 0.025
8042~ concentration increased by 1-4%/yr, while H+ has
remained unchanged.
has increased by 4-13%/yr.
and S042~ loads have increased [% slopes: 12-15% (N03~)
and 0.5-0.7% (S042~) for the Canton and Hinckley sites,
respectively] due partially to an increase in the amount of
precipitation.
Observed Characteristics
1. In the East Branch of the Sacandaga River, S042~
concentrations exceed HC03~ concentrations. USGS
monitoring of the river from 1965 to 1978 indicate an
increase in N03~ (4 peq/L.yr), a decrease in
S042~ (4 peq/L.yr), and a decrease in alkalinity (83
peq/L.yr) (Peters et al. 1981).
2. In a 1975 survey of 214 Adirondack lakes at high
elevations, pH ranged from 4.3 to 7.4. Fifty-two percent
of the lakes had pH <5.0; 7% pH 5.5-6.0 (Schofield
1976c).
3. For a subset of 40 of these 214 lakes, historic data on pH
and fish populations are available from the 1930s. Over
this period, the number of lakes with pH <5.0 increased
from 3 (out of 40) to 19. Likewise the number of lakes
without fish increased from 4 to 22. In both surveys, none
of the lakes with pH <5.0 had fish.
4. For 138 Adirondack lakes, a comparison of color-metric pH
measurements for the 1970s vs. 1930s indicated a general
decrease in pH (Pfeiffer and Festa 1980).
5. pH depressions in streams during spring snowmelt and
periods of heavy rainfall have been observed (Driscoll et
al. 1980; Galloway et al. 1980b).
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3-174
THE HUBBARD BROOK ECOSYSTEM, NEW HAMPSHIRE
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6. Based on a comparison between lakes in the Adirondack
region and within a given lake or stream monitored over
time for a one- or two-year period, elevated aluminum •
concentrations have been demonstrated to be associated with •
low pH (Driscoll et al. 1980; Schofield 1976).
7. Current status of fish populations (presence/absence) in |
Adirondack lakes and streams is clearly correlated with pH
level. The occurrence of fish is reduced particularly at •
pH levels below 5.0 (Colquhoun et al. 1980; Pfeiffer and I
Festa 1980; Schofield 1976). In the 1975 survey of 214 •
high elevation lakes, in 82% of the lakes with pH < 5.0 no
fish were collected. For lakes with pH >5.0, about 11% I
had no fish collected (Schofield 1976b). •
8. The New York Department of Environmental Conservation •
reported (based on available data) that 180 lakes have lost |
their fish populations (Pfeiffer and Festa 1980). Although
no alternative explanations for this loss of fish are _
readily apparent, historic records are not adequate to I
definitely establish acidic deposition as the cause. *
9. Survival of brook trout stocked into Adirondack waters was
inversely correlated (p < 0.01) with elevated aluminum
concentrations and low pH (Schofield and Trojnar 1980).
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The Hubbard Brook Experimental Forest (HBEF) was established in •
1955 by the United States Forest Service as the principal •
research area for the management of watersheds in New England.
The name of the area is derived from the major drainage stream B
in the valley, Hubbard Brook. Hubbard Brook flows generally B
from west to east for about 13 km until it joins with the
Pemigewasset River, which ultimately forms the Merrimack River •
and discharges into the Atlantic Ocean. Water from more than 20 I
tributaries enters Hubbard Brook along its course. Mirror Lake,
a small oligotrophic lake, discharges into Hubbard Brook at the
lower end of the valley. The HBEF is located within the White B
Mountain National Forest of north central New Hampshire. B
Although the climate varies with altitude, it is classified as
humid continental with short, cool summers and long, cold •
winters. The climate may be characterized by: (1) change- |
ability of the weather; (2) a large range in both daily and
annual temperatures; and (3) equable distribution of •
precipitation. HBEF lies in the heart of the middle latitudes I
and the majority of the air masses therefore flow from west to
east. During the winter months these are northwesterlies and
during the summer the air generally flows from the southwest. B
Therefore, the air affecting HBEF is predominantly continental. B
However, during the autumn and winter, as the colder polar air
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3-175
moves south, cyclonic disturbances periodically move up the east
coast of the United States providing an occasional source of
maritime air. The mean air temperature in July is 19°C and in
January is -9°C. A continuous snowpack develops each winter to
a depth of about 1.5m. Occasionally, mild temperatures in
midwinter partly or wholly melt the snowpack. A significant
microclimatologic feature of this area is that even the
uppermost layer of the forest soils usually remains unfrozen
during the coldest months because of the thick humus layer and a
deep snow cover.
The HBEF covers an area of 3,076 ha and ranges in altitude from
229 to 1,015 m. The experimental watershed ecosystems range in
size from 12 to 43 ha and in altitude from 500 to 800 m. These
headwater watersheds are all steep (average slope of 20-30%) and
face south. The experimental watersheds have relatively distinct
topographic divides. The height of the land surrounding each
watershed ecosystem and the area have been determined from
ground surveys and aerial photography.
The geologic substrate, outcrops of bedrock and stoney till, in
the Hubbard Brook Valley was exposed some 12,000-13,000 years
ago when the glacial ice sheet retreated northward. Bedrock is
derived from highly metamorphosed sedimentary rocks of the
Littleton formation and the granitic rocks of the Kinsman
formation. The bedrock of watersheds 1-6 is the Litleton
formation, which in this area is made of highly metamorphosed
and deformed mudstones and sandstones. It is medium to coarse
grained and consists of quartz, plagioclase, and biotite with
lesser amounts of sillimanite. Much of the area of the
experimental watersheds is covered with glacial till derived
locally from the Littleton formation. The geologic substrate is
considered watertight and losses of water by deep seepage are
minimal.
Soils are mostly well-drained spodosols (haplorthods) of sandy
loam texture, with a thick (3-15 cm) organic layer at the
surface. Most precipitation infiltrates into the soil at all
times and there is very little overland flow (Pierce 1967).
This is because the soil is very porous, the surface topography
is very rough (pit and mound, mostly from wind-thrown trees),
and normally there is little soil frost.
Soil depths are highly variable but average about 0.5 m from
surface to bedrock or till. Soil on the ridges may consist of a
thin accumulation of organic matter resting directly on the
bedrock. In some places, impermeable pan layers at depths of
about 0.6 m restrict vertical water movement and root
development. The soils are acid (pH £ 4.5) and generally
infertile.
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3-176
The vegetation of the HBEF is part of the northern hardwood
ecosystem, an extensive forest type that extends with variations
for Nova Scotia to the western Lake Superior region and
southward along the Blue Ridge Mountains. Classification of
mature forest stands as northern hardwood ecosystems rests on a
loosely defined combination of deciduous and coniferous species
that may occur as deciduous or mixed deciduous-evergreen
stands.
Deposition
(kg S042~/ha.yr) Fraction
36.4
22
38.4 + 2.5
33.7
30.0
41.6
42.0
46.7
31.2
29.3
34.6
33.0
43.4
52.8
wet
wet
bulk
bulk
bulk
bulk
bulk
bulk
bulk
bulk
bulk
bulk
bulk
bulk
Time Annual
Period Precipitation Reference
(m)
1981
1980
1964-74
Jun63-May64
Jun64-May65
Jun64-May65
Jun64-May65
Jun64-May65
Jun64-May65
Jun64-May65
Jun64-May65
Jun64-May65
1973-74
1973-74
1.50 NADP 1981-83
(Hubbard Brook)
.87 NADP 1981-83
(Hubbard Brook)
1.30 Likens et al. 1977a
Likens et al. 1977a
Observed Characteristics
1 . The external and internal generation of H"1" exerts nearly
equal roles in driving the weathering reactions. Input of
H+ is mainly in the form of H2S04 and HN03 (Likens et
al. 1977a).
2. Average streamwater pH ~ 5. During snowmelt events pH
depressions of 1.0 to 2.0 units have been reported (Likens et
al. 1977a).
3. The Hubbard Brook ecosystem accumulated hydrogen, nitrate and
ammonium ions over the period 1963-74. Over the same period
there was a net loss of SO^- (Likens et al. 1977a).
4. Ca2+ and SO^- dominated the streamwater chemistry
at the HBEF. SO^" was more than 4 times as abundant
as the next most abundant anion which was N03~ (Likens
et al. 1977a).
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3-177
5. Elevated levels of Al are found in the headwater portions of
streams in the HBEF. These levels are 2-29 times above
levels in downstream waters. This effect was attributed to
leaching of Al hydroride compounds from soils by acidic
deposition (N.M. Johnson 1979).
MAINE AND NEW ENGLAND
The 97 lakes sampled by Norton et al. (1981a) ranged in pH from
4.25 to 6.99 (median = 6.40), in elevation from 12 to 1307 m
(median = 154 m), surface area from <0.1 to 1098 ha (median =
56 ha), Secchi disc transparency from 2.5 to >5.0 m (median =
6.3 m), and water colour from 0 to 110 Pt units (median = 8 Pt
units). The bedrock of the study area was noncalcareous and
mostly granitic. As a result, the lake waters were of low
alkalinity (0-360 yeq/L, CaC03; median = 64) and specific
conductance (0-68 ymhos/cm at 25°C; median =29). The
watersheds were almost completely forested; very little cutting
had occurred in the few decades prior to sampling. Many of the
lowland lakes (fJjOO m) had cottages along their shores and
access roads in their watersheds; the high elevation lakes were
pristine and accessible only on foot. In summary, the lakes
were small to medium size, ologotrophic to mesotrophic with
moderately to very transparent water, low to moderate
concentrations of humic solutes, and low alkalinity and
conductance, and with moderately disturbed to pristine
watersheds. Haines and Akielaszek (1982) sampled a similar set
of 226 headwater lakes and streams in the other New England
states, including Maine.
Deposition
(kg S
ha.
28.
25.
24.
17.
18
36.
22
38.4+2
35
04-2-/
yr)
0
31
80
22
4
.5
Fraction
wet
wet
wet
wet
wet
wet
wet
bulk
(129.5 cm)
wet
Time
Period
1981
1981
1981
1981
1980
1981
1980
1963-74
Annual
Precipi-
tation
(m)
1.10
.87
1.15
1.10
1.50
.87
1.30
Reference
NADP
NADP
NADP
NADP
NADP
NADP
NADP
1981-83
1981-83
1981-83
1981-83
1981-83
1981-83
1981-83
Likens et al
(Acadia site)
(Bridgton site)
(Caribou site)
(Greenville
site)
(Greenville
site)
(Hubbard Brook)
(Hubbard Brook)
. 1976, 1980
(Hubbard Brook)
1981
.74
NADP
1981-83
(Bennington VI
site)
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3-178
Observed Characteristics
SUMMARY OF EMPIRICAL OBSERVATIONS
(kg S042~/ha.yr)
NORTHERN 5 wet (1980) No chemical effects
SASKATCHEWAN
17 bulk (1977)
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1. Lakewater pH declines based on comparisons with historical •
information (Davis et al. 1978) where 85% of 94 lakes studied |
(Norton et al. 1981 and 64% of 95 lakes studied (Haines and
Akielaszek 1982) were found to have lower pHs. •
2. Loss of alkalinity from lakewater in the New England states
averaging about 98 )jeq/L for 56 lakes for which there was
historical information (Haines and Akielaszek 1982). I
3. Paleolimnological confirmation of pH declines in lakes (Davis
et al. 1982). Cores from New England acidic clear water •
lakes (pH less than 5.5) with undisturbed drainage basins (5 |
of the 30 lake samples taken over at least the last 50 years)
show declines in sediment concentrations of Zn, Ca, Mg and Mn «
starting as early as about 1880 suggesting increased leaching •
of sediment delutus prior to entry into the lakes (Davis et
al. 1982; Kahl et al. 1982) or reduced sedimentation rate.
4. Accelerated cation leaching from watersheds (Kahl and Norton •
1982).
5. Lakes of pH <5 are distributed throughout a range in |
elevation from 10 to 1000 m. High elevation lakes (>600 m)
tend to have low pH and alkalinity. All but two lakes having _
pH <5.5 were also less than 20 ha in surface area. I
Alkalinity and pH also increased with stream order (Haines ^
and Akielaszek 1982). Of 226 lakes and streams sampled 25%
had alkalinity 120 Veq/L, 41% were 1100 Meq/L and 50% were •
1200 Ueq/L (Haines and Akielaszek 1982).
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SUMMARY
Location Deposition Summary Effects •
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ELA, ONTARIO 5.9 wet (1980) No effect I
9 and 11 bulk (1972-73) •
MINNESOTA 10-15 wet (NovSO) No effect •
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3-179
NORTHERN WISCONSIN
ALGOMA, ONTARIO
NOVA SCOTIA
MAINE
HUBBARD BROOK,
NEW HAMPSHIRE
MUSKOKA-HALIBURTON
16-17 wet (1981)
24.7 wet (1981)
22 wet (1981) APN
(Kejimkujik)
17 wet (1980) APN
(Kejimkujik)
22.5 wet (1977-80)
(CANSAP-Kemptville)
13.2-32 (various years)
(CANSAP - various
N.S. sites)
17-28 wet (1981)
36 wet (1981)
22 wet (1980)
33-53 bulk (1963-74)
23-29 wet (1976-78)
31-42 bulk
No effect
pH depression 2.1 pH
units
Elevated excess
sulphate relative to
region not receiving
acidic deposition
More lakes of low pH
than expected
Relationship between
fish and alkalinity
Loss of Atlantic
salmon populations
(historic record).
Historic record of
decreased pH in
river
Evidence of slight pH
decrease in lakes
(historic records)
No effects on
Atlantic salmon
No evidence of
effects on fish in
inland lakes
Spring pH depressions
No long term change
in stream or lake pH
1963-present
pH depressions
Fish kill associated
with pH depression in
one lake
Algal composition
related to pH
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3-180
LAURENTIDE PARK,
QUEBEC
22.2-40 wet (1977-80)
ADIRONDACKS
32-40 wet (1978-79)
29 wet (1980)
34-37 bulk (1965-78)
Indications of
decrease in pH in
some lakes,
especially in
southern region of
park (increases in
some lakes,
especially along
roads); indication of
decline in angling
success in lower pH
lakes; pH depression
in lakes in spring
(Moreau et al. 1982)
and lower pH in lakes
in spring than summer
(Bob€e et al. 1982)
Evidence of pH
declines and loss of
fish populations over
time
Detailed studies of watersheds have been carried out in
sensitive regions of North America and Scandinavia under a range
of sulphate deposition rates. The results of watershed studies
conducted in North America are described below.
For those regions currently experiencing sulphate in
precipitation loading rates of ^17 kg/ha.yr there have been no
observed detrimental chemical or biological effects.
For regions currently experiencing between 20 and 30 kg/ha.yr
sulphate in precipitation there is evidence of chemical
alteration and acidification. In Nova Scotia rivers, historical
records of salmon population reductions as documented by 40
years of catch records have occurred as well as reductions in
stream pH. In Maine there is evidence of pH declines over time
and loss of alkalinity from surface waters. In Muskoka-
Haliburton there is historical evidence of loss of alkalinity
for one lake. There is documentation of pH depressions in a
number of lakes and streams. Fish kills were observed during
spring melt in one lake. In the Algoma region there are
elevated sulphate and aluminum levels in some headwater lakes.
For regions currently experiencing loading ^30 kg/ha.yr there
are documented long-term chemical and/or biological effects and
short-term chemical effects in sensitive surface waters.
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3-181
In Quebec, sulphate concentrations in surface waters decrease
towards the east and north in parallel with the deposition
pattern. Sulphate concentrations are equal to or greater than
the bicarbonate concentration in lakes in the south west part of
the Province.
In the Adirondack Mountains of New York, comparison of data from
the 1930s with recent surveys has shown that more lakes have
been acidified. Fish populations have been lost from 180 lakes.
Elevated aluminum concentrations in surface waters have been
associated with low pH and survival of stocked trout is reduced
by the almuninum.
In the Hubbard Brook study area in New Hampshire there is pH
depression in streams during snowmelt of 1 to 2 units. Elevated
levels of aluminum were observed in headwater streams.
3.9.2.2 Short-Term or Episodic Effects
While current and historical survey data may imply long-term trends,
the samples usually represent only one or a few measurements at
any one location and are usually collected only during the summer.
This limited sampling period provides no record of pH and other
chemical changes which take place in relation to seasonal cycles or
major weather events. If short-term changes in water chemistry
coincide with sensitive periods in the life cycle of fish,
significant mortality and reduced reproduction can occur.
Severe pH depressions in streams and small lakes due to snowmelt have
been documented in a number of locales (e.g., Kahl and Norton 1982;
Schofield 1973). The depression may be as much as 1-2 pH units.
Much of the metal content of the snowpack is also released in early
melting stages. Thus critical hydrogen ion and trace metal levels
may be reached temporarily, even in waters with relatively high
summer pH values. Leaching of metals from soils and sediments may be
especially severe during this period, resulting in pulses of high
concentrations of potentially toxic metals (e.g., Al > several
hundred ppb [Kahl and Norton 1982; Schofield and Trojnar 1980]).
The question has often been raised, "How long does it take before the
lakes become acidic?" The previous sections have indicated relation-
ships for lakes and streams which have already been acidified.
However, the rate of change is one of the least well-defined aspects
of the acidification phenomenon. The rate-of-change questions become
less relevant in light of evidence that current acid loadings are
causing damage to fisheries and other biota due to short-term
exposures to low pH and associated high metal concentrations, as
reviewed earlier in this report.
The pH of lakes or streams tends to fluctuate considerably during the
year, and average annual pH is a composite of these patterns. Thus,
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3-182
I
organisms which may respond to extreme concentrations of H+ or
metals over a few days. This, plus the known significance of brief •
acute exposure (Spry et al. 1981), suggests that the magnitude and I
duration of short-term increases in H+, associated with a defined
"flushing event", could be used for further evaluating critical dose/
response relationships in stream ecosystems, and lakes. •
Research on brook trout and white sucker by Baker (1981), Baker and
Schofield (1980), on Atlantic salmon by Daye (1980) and Daye and •
Garside (1975, 1977, 1980), and related research by Beamish and J
Harvey (1972), Beamish (1974a, 1974b, 1976), and Harvey (1975, 1979,
1980) has provided a broad understanding of the response of several _
pH-sensitive fish species to both long-term and short-term elevated •
H and aluminum exposures. Mortalities have been documented for
chronic pH depression, and effects on egg viability, hatching success
and adult survival for short-interval acute H+ and aluminum •
exposures are reasonably well known (Baker and Schofield 1980). I
Among the experimentally-based relationships developed by Daye, •
Garside, Baker and Schofield is a recurring pattern (Loucks et al. •
1981):
1) the short-term acute exposure, or "shock", effects, •
including responses to aluminum, can take place in two to ™
four days of exposure, with as little change as 0.5 to 1.5
units of the pH scale; and H
2) these shock exposures can be expected to occur in waters
with a broad range of pH above the level at which chronic •
effects occur. •
Stream water chemistry studies from a number of locations
(Table 3-26) show short-term pH depressions during snowmelt and storm •
events (e.g., 1.0 unit on the Shavers Fork River in West Virginia B
[Dunshie 1979]) and from 1.0 to 2.0 units in two watersheds being
studied in the Adirondacks (Galloway et al. 1980b). A third lake •
studied by Galloway et al. (1980b) at the Adirondack site had a mean |
annual pH of about 4.8 and shows no pH depression during flushing
event. Likens et al. (1977a) reported pH depressions of 1.0 to 2.0 ^
units for Hubbard Brook, New Hampshire. Outside the regions with •
snow accumulation, the maximum pH decline during a flushing event
appears to occur during major rainfall events following a rain free
period (Dunshie 1979). fl
Sulphate loadings associated with observed short-term pH declines and
resulting biological effects are summarized in Table 3-26. In the •
ELA, Ontario, annual loadings of sulphate in precipitation of about |
10 kg S042~/ha.yr have generally resulted in pH declines of
only 0.2-0.3 units and no apparent biological effects. Depressions _
in pH of 0.3-1.0 units have been observed in northern Minnesota •
streams receiving approximately 14 kg S042-/ha.yr. However,
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3-183
TABLE 3-26. PERIODIC pH DEPRESSIONS OBSERVED IN STREAMS AND LAKES WITH DIFFERENT SULPHATE LOADINGS AND
CORRESPONDING BIOLOGICAL EFFECTS. SURFACE WATER ALKALINITIES IN THESE AREAS ARE GENERALLY
LESS THAN 200 yEQ/L.
Location
(kg
Annual
Sulphate
Loading
S042~/ha.yr
Lowest
PH
Observed
Largest
Between
pH
Difference
Spring pH
Observed Biological Effect
and summer or
by wet deposition)
w i nter
val ues
Tovdal R. Norway
L. Timmevatten Sweden
(1970)
Sweden (1972)
Hubbard Brook
Experimental Forest
Panther L. ILWAS
Project New York
40
40
40
pH shock suspected but no field
measurements taken during the
fish kill
4.2
4.3
0.8
1.1
Harvey et a 1. 1982
Mills pers. comm.
Keller and Gale 1982
Siege) 1981
Church and Galloway 1983
Fish kilI (sea trout)3
Wild population of minnows
have disappeared*3
Caged sea trout and minnows
experienced 68$ and 59%
mortality6
No biological studies
No biological data available;
fish population 1st from one
lakeJ
Muskoka-Hal i burton 30 4.1 1.1
Ontario (4 streams)
( lake outflows) 30 4.8 1.3
Plastic Lake 30 4.0 1 .7
Ontario Inlet
Outlet 30 5.0 0.7
Shavers Fork W. Virginia 30 5.6 0.9
(stream)
Algoma 5.0 2.1
Fi Ison Creek, 14 5.5 0.3-1
Northern Minnesota
Experimental Lakes 10 4.5 has been 0.2-0.3
Area Ontario observed generally
above 5
Evidence of fish population
damage in areas lakes0 and
actual algae species^
100? mortality of caged
rainbow trout'
13? mortality of caged
rainbow trout^
Conditions caused by heavy
rain; no biological studies6
No biological studies11
No biological studies'
No apparent biological
effects
a Braekke 1976
b Hultberg 1977
c Harvey 1980
d Nichol Is et al . 1981
e riiinchia 1Q7Q
-------
3-184
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the lowest pH reading recorded is 5.8 and no biological studies have
been conducted.
Galloway and Dillon (1982) have examined the relative importance of •
sulphuric and nitric acids in causing alkalinity (and pH) reduction
during snowmelt and conclude that a major portion of the reduction in
alkalinity during snowmelt was attributable to nitric acid. Although •
•
itself showed little variation during snowmelt, its
continued large presence in the stream was responsible for the
alkalinity reduction in an indirect manner, namely by causing •
long-term alkalinity reductions (as opposed to episodic). Thus, the |
episodic reduction of alkalinity due to NOg" is added to the
long-term reduction in alkalinity due to SO^". Jeffries et jm
al. (1981) demonstrated that in Muskoka-Haliburton the increase in •
hydrogen ion concentration in several streams during snowmelt was due
to increases in both N(>3~ and SO^".
3.9.2.3 Sensitivity Mapping and Extrapolation to Other Areas of
Eastern Canada J|
TERRESTRIAL
In order to identify the magnitude of the surface water acidification •
problem our ability to extrapolate the results of the detailed
watershed study areas to the remainder of eastern North America must
be determined. Within eastern North America are hundreds of •
thousands of lakes and streams and it is clearly impractical to •
establish detailed or regional hydrochemical monitoring for them all.
However, there is an urgent need to determine if the watershed study fij
areas currently being monitored are anomalous in terms of their |
geochemical characteristics or if, in fact, they are representative
of conditions occurring over large areas of eastern North America.
An early approach to this problem in Canada was to consider all of ™
the Precambrian Shield as "sensitive" and then assume any study area
located anywhere on the Shield would be representative of over 75% of •
eastern Canada (Altshuller and McBean 1979). This approach implied |
that the Canadian Shield was a single granitic plate and not, as is
the case, a number of complex geological provinces composed of a «m
variety of rock types (including marble) and covered, in places, by •
unconsolidated material of varying texture and carbonate content
(Section 3.5). Areas outside the Shield, where hydrochemical changes
have occurred (e.g., the Maritime Provinces), also exhibit a range of 9
soil and bedrock conditions. ™
A major drawback to more detailed analyses and extrapolation has been •
the lack of the analyses of information on surficial and bedrock j|
geological conditions for all of eastern North America, in a regional
but detailed form. This has recently been alleviated for Canada with M
bedrock sensitivity mapping of Shilts et al. (1981) which has been •
incorporated into the bedrock-soil mapping composite of Lucas and
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3-185
Cowell (1982). This mapping is discussed in more detail in
Section 3.5. In order to determine the representativeness of three
of the detailed watershed study areas, (Algoma, Muskoka-Haliburton
and southwest Nova Scotia), the 65 classes of soil and bedrock
characteristics mapped by Lucas and Cowell (1982) will be utilized.
The basis for extrapolation is the 1:1,000,000 scale map shown in
Figure 3-9 (in map folio). This mapping represents the most detailed
compendium of soil and bedrock characteristics yet assembled for all
of eastern Canada. Extrapolation has been carried out by reviewing
the kinds of soil and bedrock terrains which form the geochemical
templates of three of the watershed study areas and then determining
how representative these areas are in eastern Canada.
The 65 classes of terrain characteristics are listed in Tables 3-27
and 3-28. Each class is identified according to a two or three
character alpha-numeric code which is defined in Table 3-8.
Table 3-27 lists the area and percent cover of each class north and
south of 52°N latitude for each province. [Figure 3-9 shows only the
areas south of 52°N.] Table 3-28 summarizes the area and percent
cover of each class for all of eastern Canada. This table indicates
that 54% of eastern Canada is composed of bedrock types in
combination with soil types which have a low potential to reduce
acidity. These are predominately noncalcareous sands and sandy tills
overlaying granitic-type bedrock. Within the area south of 52°N, 51%
or 911,089 km^ is considered as having a low potential to reduce
acidity of atmospheric deposition prior to entering surface waters.
Terrain Characteristics of Three Specific Study Areas
Terrain classes are based on bedrock geology, percent bedrock
exposed, soil depth and soil texture or depth to carbonate
(Table 3-8; Section 3.5). Table 3-29 shows the terrain classes for
watersheds within which the study lakes and rivers occur. These
results have been obtained by directly overlaying the watershed areas
for Algoma, Muskoka-Haliburton and Nova Scotia onto Figure 3-9.
By far the greatest proportion of each area is composed of terrain
classes interpreted as having a low potential to reduce the acidity
of atmospheric deposition (69 to 98%). The most complex area and the
one with the greatest range of terrain conditions is Algoma which has
up to 69% has a low potential to reduce acidity, 25% interpreted as
having a moderate potential, almost 5% with a high potential to
reduce acidity and less than 1% organic terrain (which has not been
interpreted).
Two low potential terrain classes dominate in each area. In Algoma
and Muskoka-Haliburton these are the L3 (41.79% and 59.42%,
respectively) and the L4c (21.05% and 32.25%, respectively) classes;
in Nova Scotia these are the L4b (53.15%) and L4c (27.11%) classes.
-------
3-186
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3-187
TABLE 3-28. SUMMARY OF TERRAIN TYPES AND POTENTIAL TO REDUCE ACIDITY FOR ALL OF EASTERN
CANADA
Terrain Types
(Potential to
Reduce Acidity)
Hla
Hlb
Hie
Hie
Hlf
Hlg
H1h
Hli
Hlj
H2a
H2b
H3a
H3b
H3c
Total
High
Potential
Mia
Mic
Mid
Mlf
Ml i
Mlj
Mln
Mlo
Mlp
Mlq
Mir
Mis
Mlt
South of
(area =
km2
43,632
65,690
8,105
1,004
109
7,989
5,305
8,959
1,405
1,283
3,237
15,933
61,555
83,914
308,390
86
20
174
48
3,956
698
52'N Latitude
1,779,436 km2)
% of Zone
2.45
3.71
0.46
0.06
0.00
0.45
0.30
0.50
0.08
0.07
0.18
0.90
3.46
4.72
17.33
<0.01
<0.01
0.01
<0.01
0.22
0.04
North of
(area = 1
km2
35,258
3,034
2,230
4,634
3,915
17,506
66,577
7,018
83
5,523
9,615
2,325
489
1,038
3,114
566
10,389
374
1,520
52 °N Latitude
,357,595 km2)
% of Zone
2.60
0.22
0.16
0.34
0.29
1.29
4.90
0.52
0.01
0.41
0.71
0.17
0.04
0.08
0.23
0.04
0.77
0.03
0.11
Total for
(area =
km2
78,890
65,960
8,105
1,004
109
7,989
5,305
8,959
1,405
4,317
5,467
20,567
65,470
101,420
374,967
7,104
83
5,543
9,615
2,325
489
1,038
3,114
740
48
14,345
374
2,218
Eastern Canada
3,137,031 km2)
% of
Eastern Canada
2.51
2.10
0.26
0.03
< 0.01
0.25
0.17
0.29
0.08
0.14
0.17
0.66
2.09
3.23
11.95
0.23
< 0.01
0.18
0.31
0.07
0.02
0.03
0.10
0.02
<0.01
0.46
0.01
0.07
-------
3-188
TABLE 3-28. CONTINUED
Terrain Types
(Potential to
Reduce Acidity)
M1u
Mlv
M2a
M2b
M3
M4a
M4b
M4c
M5
M6a
M6b
M7a
M7b
M7c
Total
Moderate
Potential
Lib
Lie
Lid
Lie
L2a
L2b
L3
L4a
L4b
L4c
L4d
Total
Low
Potential
South of 52'N Latitude North of 52'N Latitude Total for Eastern Canada
(area = 1,779,436 km2) (area = 1,357,595 km2) (area = 3,137,031 km2)
km2
82
982
13,662
1,564
117,987
9,382
7,749
10,104
32,237
18,023
83,473
46,831
345,058
143
5,064
914
705
2,110
369,467
11,226
109,262
386,090
21
911,089
% of Zone km2
<0.01
0.06
0.77
0.09
6.63
0.53
0.44
0.57
1.81
1.01
4.58
2.63
19.39
0.01
0.28
0.05
0.04
0.12
20.76
0.63
6.14
21.70
< 0.01
51,20
415
14
46,726
13,670
84,180
6,027
4,051
19,060
17,523
22,933
17,973
274,626
3,322
6,395
75,736
9,491
46,755
40
157,723
4,369
52,247
290,162
788,920
% of
% of Zone km Eastern Canada
0.03
0.01
3.44
1.01
6.20
0.44
0.30
1.40
1.29
1.69
1.32
20.23
0.24
0.47
5.58
0.70
3.44
<0.01
11.62
0.32
3.85
21.37
58.11
415
14
82
982
60,388
15,234
202, 167
9,382
13,776
14,155
51,297
35,546
104,406
64,804
619,684
3,322
6,538
80,800
10,405
47,460
2,150
527,190
15,595
161,509
676,252
21
1,700,009
0.01
< 0.01
< 0.01
0.03
1.93
0.49
6.44
0.30
0.44
0.45
1.64
1.13
3.33
2.07
19.76
0.11
0.21
2.58
0.33
1.51
0.07
16.80
0.50
5.15
21.55
<0.01
54.19
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TABLE 3-28. CONTINUED
Terrain Types
(Potential to
Reduce Acidity)
Ola
Olb
Olc
Old
02a
02c
02d
03a
03c
03d
Total
Organic
Terrain
South of
(area =
km2
51,349
15,799
40,598
106,519
34
170
48
55
327
214,899
52'N Latitude
1,779,436 km2)
% of Zone
2.89
0.89
2.28
5.99
<0.01
0.01
<0.01
<0.01
0.02
12.08
North of
(area =
km2
154,399
24,627
12,351
35,681
207
207
227,472
52*N Latitude
1,357,595 km2)
% of Zone
11.37
1.81
0.91
2.63
0.02
0.02
16.76
Total for
(area =
km2
205,748
40,426
52,949
142,200
34
207
377
48
55
327
442,371
Eastern Canada
3,137,031 km2)
% of
Eastern Canada
6.56
1.29
1.70
4.53
< 0.01
< 0.01
0.01
0.01
< 0.01
0.01
14.10
-------
TABLE 3-29.
Terrain
Class
L2b
L3
L4a
L4b
L4c
Total L
M4b
M7a
M7b
M7c
Total M
Hlb
Hlc
Hli
Total H
Ola
Olc
Old
Total 0
Study Area
TERRAIN CHARACTERISTICS OF WATERSHEDS
STUDY AREAS OF EASTERN CANADA
Algoma Muskoka-Haliburton
km2 % km2 %
116.1 6.52
3,380.7 41.79 1,058.1 59.42
283.9 3.51
225.8 2.79
1,703.2 21.05 574.2 32.25
5,593.6 69.14 1,748.4 98.19
1,838.7 22.73
109.7 1.36
25.8 0.32
116.1 1.44 19.4 1.09
2,090.3 25.85 19.4 1.09
45.2 0.56
264.5 3.27
77.4 0.96
387.1 4.79
12.9 0.16
6.5 0.08 12.9 0.72
19.4 0.24 12.9 0.72
8,090.4 1,780.7
3-190
CONTAINING THE DETAILED
Southwest Nova Scotia
km2 %
154.8 1.36
6,045.2 53.15
3,083.9 27.11
9,283.9 81.62
1,419.4 12.48
1,419.4 12.48
509.7 4.48
161.3 1.42
671.0 5.90
11,374.3
1
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3-191
It is assumed that these classes represent the terrestrial
geochemical template for the three study areas. The other "low"
potential classes are very minor in these watersheds and one would
expect little or no effect of acidic deposition in basins dominated
by "moderate" and "high" potential templates.
In the Muskoka-Haliburton watersheds, nine of the lakes and
associated tributary streams which have been monitored closely occur
entirely within the L3 class. Detailed lake basin mapping by
Jeffries and Snyder (1983) for 6 of the lakes indicate that this L3
class is predominately composed of their "Minor Till Plain" and "Thin
Till" classes overlaying gneiss bedrock. These two surficial types
represent between 84.3 and 94.0% of the basins of Red Chalk, Blue
Chalk, Chub, Dickie, Harp and Jerry lakes.
The three dominant terrain classes in these study areas (L3, L4b and
L4c) are composed of the following: (1) L3 - shallow sands and
acidic type rocks (granite, gneiss, quartzite or other alkalic rocks)
which outcrop in 0-49% of the map area; (2) L4b - deep sands
overlaying ultramafic, serpentine and noncalcareous sedimentary rocks
outcropping in 0-49% of the unit; and (3) L4c - deep sands
overlaying bedrock similar to L3. These classes represent dominant
conditions in a map area. At this scale of mapping (1:1,000,000)
other subdominant conditions probabl} occur. However, the evidence
from more detailed mapping at Muskoka-Haliburton, as described above,
indicates that the descriptions are representative. It should be
noted further that the term "sands" refers to the matrix texture; the
deposit it represents is most commonly a till or glacial-fluvial
outwash which include larger sized fragments.
Results of Terrain Extrapolation
Table 3-27 provides the basis of extrapolation by province and
Table 3-28 for all of eastern Canada. Terrain classes L3, L4b and
L4c, which represent the major geochemical templates for the
watershed study areas, are three of the four most common terrestrial
types. In eastern Canada, they cover 17% (527,190 km2), 5%
(161,509 km2) and 22% (676,252 km2) respectively (Table 3-28).
They represent over 80% of the sensitive terrain types in Eastern
Canada. Other classes which cover significant areas but are not
represented in the study areas are H3c (deep clay overlying granitic
rocks), Ola (organic deposits overlying carbonate rocks), Old
(organic deposits overlying granitic rocks), and L2d (shallow sand
overlying granitic rocks with 50-74% outcropping).
Approximately one-fifth of eastern Canada (690,117 km2) currently
receives loadings of about 20 kg/ha.yr or more of SO^2" in
precipitation in 1980. Within this loading zone terrain classes L3,
L4b and L4c cover 18% (127,237 km2), 6% (40,222 km2) and 22%
(153,545 km2) respectively. In total, the three terrain types
cover 46.52% of eastern Canada within the 20 kg/ha.yr, or higher,
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3-192
AQUATIC
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loading of SO in precipitation. This is an area of
321,004 km2 (125,192 mi2) which represents 99% of all those areas
with the lowest potential to reduce acidity within this loading zone. fl
These areas occur primarily on the Grenville Province of the I
Precambrian Shield in southern Quebec and Ontario as well as in the
Appalachian Region of New Brunswick and Nova Scotia (Figure 3-9). •
These results indicate that over one-half of eastern Canada, is
representative of terrain characteristics (Table 3-8) under which
aquatic acidification effects have been observed. •
From these results, it is concluded that terrain characteristics
in the three watershed study areas are correlated with measured ft
acidification effects, especially as expressed by alkalinity 0
measurements. These three study areas are not anomalous but are
representative of larger portions of Eastern Canada as defined by mm
these terrain characteristics. •
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As shown in the previous section the bedrock and surficial geology of
the study areas are typical of large areas of eastern Canada.
However specific watersheds with varying glacial deposits (kame,
spillway, till, etc.) rock component hardness (i.e., resistance to ^
weathering) and varying hydrological characteristics result in W
surface waters of varying alkalinity and total cation concentrations
within each study area.
Hydrochemical data from the Muskoka-Haliburton area of Ontario also ml
compares closely with mapped terrain conditions. Average annual and
spring T.I.P. alkalinity values for 9 lakes within the Muskoka- •
Haliburton study area are all lower than 71yeq/L (Table 3-30). Five jj
of these lakes are considered very sensitive on the basis of their
alkalinity regime (<40 peq/L). The basins of all 9 lakes are _
composed primarily of shallow to deep (<2 m) sandy tills overlaying •
gneiss (class L3 and L4c). In addition there is a close correlation
between terrain class and alkalinity regime for a population of 141
lakes sampled throughout Haliburton County and Muskoka District. M
Table 3-31 shows the occurrences of lake alkalinities grouped by v
sensitivity classes, in each of the mapped terrain types. There is
clearly a strong relationship with 77.5% of the lowest alkalinity J|
lakes (0-39.9 and 40-199.9 yeq/L) occurring in terrain classes L3 and J
L4c. It is not possible, at present, to extrapolate the results of
Table 3-31 to all the areas of eastern Canada mapped in these two —
terrain classes. I
Further support for the representativeness of the study areas is
drawn from the water quality data. Figures 3-49 and 3-50 show the 4|
distribution of lake alkalinities for a series of geographical areas •
on sensitive and moderately sensitive terrain. The data are taken
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3-193
TABLE 3-30. AVERAGE ANNUAL OR SPRING TOTAL INFLECTION POINT
ALKALINITIES FOR NINE LAKES IN THE MUSKOKA-HALIBURTON
WATERSHED STUDY AREA (data from Ontario Ministry of
Environment)
Lake
Time of Record
Alkalinity
mg/L yeq/L
Harp
Dickie
Chub
Red Chalk
Blue Chalk
Jerry
Plastic
Heney
Crosson
1979-80
1979-80
1979-80
1979-80
1979-80
1979-80
Spring/79
Spring/79
Spring/80
3.32
0.762
0.798
3.15
3.53
3.31
0.62 +; 0.5
0.34 +_ 0.5
0.49 + 0.5
66.4
15.24
35.96
63.0
70.6
66.2
12.4 H- 10.0
6.4 + 10.0
9.8 + 10.0
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TABLE 3-31.
3-194
DISTRIBUTION OF 141 LAKE ALKALINITIES, GROUP BY
SENSITIVITY CLASSES, IN VARIOUS TERRAIN TYPES OCCURRING
IN HALIBURTON COUNTY AND MUSKOKA DISTRICT, ONTARIO
Terrain
Class
L3
L4C
L2d
L2b
Hlc
Hli
M4b
M7c
Old
Map
Area
(km2)
4283.9
1645.2
206.5
141.9
109.7
51.6
25.8
45.2
25.8
Alkalinity
0-39.9 40-199.9
34 (24.1) 61 (43.4)
5 (3.5) 9 (6.4)
3 (2.1) 3 (2.1)
4 (2.8) 3 (2.1)
2 (1.4)
Classes (yeq/L)
200-499.9 500
3 (2.1) 6 (4.4)
7 (4.9)
1 (0.7)
Total
6535.6 46 (32.5)
78 (55.4) 11 (7.7)
6 (4.4)
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M
_ra
O
>>
+••
'c
O
03
0
0)
JC
ca
100
80
60
40
20
0-
100
80
60
40-
20
0
100
80
60
40
20
0
100
80
60
40
20
0
100
80
60
40
20
0
100
80
60
40
20
0
<0 ^0-39^^40-199 200-499
>500
3-195
BRUCE AND GREY COUNTIES
n=10
<0 ' 0-39.9 40-199 200-4991 >500
<0 ' 0-39.9 40-199 200-499 >500
<0 0-39.9 40-199 200-499 >500
<0 '0-39.9 40-199 200-4991 >500
HALIBURTON COUNTY
n=197
MUSKOKA DISTRICT
n=159
KENORA DISTRICT
(S. of 51° Lat.)
RAINY RIVER DISTRICT
n=99
Figure 3-49.
ALGOMA DISTRICT
n=449
<0 0-39.9 40-199 200-499 >500
Alkalinity (peq/L)
Distribution of alkalinity values for lakes in six
regions on Ontario.
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3-196
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3-197
from Table 3-12. The percentage distribution of lake alkalinities
are similar in all areas and contrast strongly with the alkalinities
of 10 lakes in Bruce and Grey Counties which are located on
calcareous till in southern Ontario (nonsensitive terrain).
While the alkalinity distributions are similar, there are some
important differences. The distributions for Haliburton, Muskoka,
and Algoma have already been altered in that there is a greater
number of lakes with low alkalinity than in the Kenora or Rainy River
Districts or in background areas such as Northern Saskatchewan.
Dillon (1982) further demonstrated the differences in alkalinity
values for lakes in the areas of higher sulphate deposition (Muskoka-
Haliburton and Parry Sound) by plotting the cumulative distributions
(Figure 3-50). It is accepted that alkalinity distributions are
already influenced by acid loadings in some areas and to reflect
natural conditions the distributions should be shifted to the right
as plotted in Figures 3-49.
Within each study area, the number of lakes for which detailed data
are available is small relative to the total number of lakes.
Therefore, it is important to show that the intensive study lakes and
rivers themselves are representative of the surface waters of the
sensitive areas. There are a total of 18 calibrated study lakes at
ELA (1), Algoma (5), Muskoka-Haliburton (8), Quebec (1) and Nova
Scotia (3). The current alkalinities show 2 less than 0 peq/L, 7 in
the 0-40 yeq/L range and 9 in the 40-200 yeq/L range. Lakes above
200 are not subjected to intensive studies since acidification
effects are minimal. In addition, Ontario has extensive information
on five calibrated lake studies near the point sources in Sudbury
which is used to contrast effects of local sources and long range
transport. Of the 22 rivers in Nova Scotia used in analysis of
salmon catch data, current alkalinities range from less than zero
(acidic) to 173 yeq/L (Figure 3-47).
The study lakes and streams are located in areas with terrain
characteristics and have alkalinity values similar to other sensitive
areas in Canada. Therefore, the effects observed in the study lakes
and rivers in response to specific loading rates should be similar in
other water bodies in these sensitive areas. Similarly, loading
rates protective of these study lakes should be protective of other
sensitive waterbodies.
Possible Magnitude of Effects
The Canadian members of the Work Group have concluded that an
indication of the extent of the current water quality effects may be
derived for all of Ontario using the information presented in
Section 3.6.1. The Precambrian area east of Algoma contains some
50,000 lakes (Cox 1978). The distribution of alkalinity values for
lakes in districts within the 20 kg/ha.yr wet SO^~ deposition
isopleth (from Table 3-12) indicates that about 20% or about 10,000
lakes have alkalinity values and acid loadings that are combining to
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3-198
3.9.3 Use of Acidification Models
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currently cause pH depression to values (less than 5.5) likely to be
causing biological damage.
Cox (1978) has indicated that the lake counts underestimate the •
number of lakes with surface areas less than one hectare by as much
as a factor of three so the 50,000 and 10,000 numbers are both
underestimates.
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The data for the 57 headwater streams in Muskoka-Haliburton show that
65% experience minimum pH values less than 5.5 and 26% have minimum
pH values less than 4.5 (Figure 3-21). Although the total number of
miles of streams within the 20 kg/ha.yr wet SO^- deposition
isopleth is not known and quantitative extrapolations are not «
possible, it is clear that many miles of stream water must also •
currently experience pH depressions to levels that can potentially
cause biological damage.
There is a larger area of lakes underlain by Precambrian rock in V
Quebec and the Maritime provinces where the acid loadings are at
least as much as those at Algoma. While specific lake count data are •
not available, it is likely that many thousands of lakes are ||
currently receiving acidic deposition.
In both Ontario and Quebec many more thousands of lakes are slightly •
less sensitive to acidic deposition and may experience biological *
damage in the future if the acid deposition continues.
_
Precambrian areas of eastern North America is measured in the tens of
thousands with even more sensitive to effects in the future. •
The U.S. members of the Work Group believe the statements in this
section cannot be supported by the facts. The combined analysis of _
lake survey data, terrestrial mapping data and deposition data is an •
interesting methodology. Pending validation, the U.S. members have ™
too many concerns about the influence of uncontrolled variables to
consider its use more than speculative. One important variable is I
the level of dry deposition from local sources which can affect the •
representativeness of the survey lakes. Other factors which may
determine the overall neutralizing capacity of a watershed system in •
addition to terrain class include elevation, hydrologic routing time, \l
lake morphometry and vegetative cover. We therefore cannot support
the conclusions in this section in the absence of further _
methodological validation. •
I
A number of process-oriented (mechanistic) models have been developed
(or are under active development) that simulate in detail the flow •
of acidic precipitation through terrestrial systems and the resulting •
chemical response of surface waters. These models have the potential
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3-199
to predict stream and lake responses (e.g., pH depressions) to
episodic events, but most of them are not suitable for predictions of
long-term ecosystem responses. Examples of these process-oriented
models include the ILWAS model (Chen et al. 1982), the Birkenes model
(Christophersen et al. 1982), and the trickel-down model (Schnoor
et al. 1982). Each of these models has achieved some success in
relating short-term variations in water chemistry of small drainage
basins to hydrology and chemistry of precipitation. These models,
while calibrated for specific watersheds have not been validated on a
temporal or spatial scale that permits their general application with
significant confidence.
More global modeling efforts, such as those of Hough et al. (1982),
and USFWS (1982) have formulated detailed mechanistic submodels but
have not developed them to the level of working codes. Thus,
prediction of the dynamic response at the aquatic regime to the
atmospheric loading remains to be achieved at this time. However,
several efforts towards development of empirical or semi-empirical
steady state models relating aquatic chemistry to the atmospheric
loading stress have advanced to the point that response estimates are
possible within the limits of assumptions of the models.
Three important general points must be made about these models.
First, validation (especially for surface waters in North America)
remains to be achieved. Second, each of these models is based upon
individual, specific sets of asumptions regarding their application.
Application of these models is therefore limited by the degree to
which these assumptions are met. Third, these models are not dynamic
and therefore, determination of the rates of reaction between
sulphate deposition and lake water pH based on the models is not
possible. The models rely on steady state conditions. With these
important points in mind, potential use of these models for
quantitative estimates of the relationship of SO^- deposition
to lake pH is discussed below.
The earliest empirical acidification model was developed by Aimer
et al. (1978) and modified by Dickson (1980, 1982), who related lake
pH and excess SO^" load (concentration of excess SO^" multiplied by
annual runoff) for arbitrary classifications or groupings of Swedish
lakes. Since this relationship is, in effect incorporated by
Henriksen (1979, 1980, 1982) in his model, it will not be discussed
in detail here.
3.9.3.1 The "Predictor Nomograph" of Henriksen
Henriksen (1979, 1980) has studied atmospheric and edaphic influences
on the chemistry of oligotrophic lakes in Scandinavia and has
developed empirical formulations relating these influences to
acidification. He has derived an acidification "indicator," a
quantitative acidification "estimator," and an acidification
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3-200
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"predictor nomograph" (Henriksen 1979, 1980). Of these formulations,
only the "predictor nomograph" is intended for use as a predictive
tool. •
Henriksen (1980) developed his "predictor nomograph" of freshwater
acidification based on the hypothesis that "acidified waters are the M
result of a large scale acid base titration." He compared the •
concentration of "Ca* + Mg*" with lakewater 804* concentrations
(* indicates "excess concentration" — that above contributions from
seasalts) in the pH range 5.2-5.4 and 4.6-4.8 using data from 719 •
lakes in southern Norway (Wright and Snekvik 1978) and obtained ™
"highly significant" linear correlations. The line for the pH range
5.2-5.4 agreed very well with a theoretical titration nomograph of l|
bicarbonate concentration vs. (H+ added), assuming that bicarbonate j|
concentration is directly proportional to (Ca* + Mg*) concentration
and that (H+ added) is proportional to 804* concentration. The _
line for the pH range 4.6-4.8 did not agree with such a theoretical •
bicarbonate titration nomograph, but Henriksen (1980) argued that his
deviation was readily explained by the effects of dissolved aluminum
leached from soils. To complete his predictor nomograph, Henriksen I
(1980) added a Ca* concentration axis parallel to the (Ca* + Mg*) I
axis and a precipitation pH axis parallel to the 804* axis
(Figure 3-51). The former was derived from correlations of Ca* •
concentrations and (Ca* + Mg*) in lake waters; the latter was derived •
by combining: (1) a correlation of 864* concentration in lake water
to 804* concentration in precipitation, and (2) a correlation of ^
864* concentration in precipitation to H+ concentration in V
precipitation. Henriksen (1982) added an axis of 864* in •
precipitation parallel to the axis of 804* in lakewater based on
his 1980 regression. •
Henriksen (1979, 1980) derived his predictor nomograph for pristine,
oligotrophic lakes in areas with granitic or siliceous bedrock types •
and thin podsolic soils. In these lakes that have been receiving •
acidic deposition, 864^" is the major anion. Prior to the
advent of acidic deposition, Ca2+ and HC03~ were the dominant _
ions in these lakes. Lakes used to develop the relationships had low fl
concentration of organic acids. The lakes ranged in area from '
0.1 to 30 km2 and in 90% of the lakes the Ca+2 concentration
was less than 80 yeq/L. None of the lakes was on a major river H
(i.e., had very large watersheds) (Wright and Snekvik 1978). I
Henriksen (1980) verified the predictor nomograph with an independent •
data set from an October 1974 survey of 155 Norwegian lakes (Wright •
and Henriksen 1978). These lakes ranged in area from 0.25 to
1.0 km2, occurred at the head of undisturbed watershed drainage
basins, and constituted 5% or more of their watersheds (Wright and •
Henriksen 1978). Henriksen (1980) found that for over 85% of the ™
lakes, the nomograph correctly predicted a pH "grouping" (pH<4.7 —
"acid lakes", 4.7 pH <5.3 — "transition lakes", pH>5.3 — •
"bicarbonate lakes"). He also found that the nomograph was valid for |
18 "large lakes" in southern Norway.
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| 3-201
1
1
• 300-
™ § 200-
3-
^H **
I *<»
+
§*«
0
_ 100-
1
vv
1
0J
250-
200-
3 150-
"17
CD
3-
*«"
0 100-
50-
0-
&''
X
X
X
X
/ 1
X ft,.1 x
HCOo- Lakes xx ^xx
O ' Vx"
X ^^
X^ x-
^/^ ^^
.X ^x^
/ ^^ Acid Lakes
>x ^^
//^
////^
1 1 1 1 1 1 1 1 1 1 1
0 100 200
• SO4* in Lakewater, fyieq/L)
i i i i i i i i i i i i i i
10 50 100
SO4* in Precipitation, fyieq/L)
1
i i i i i i i i
7.0 5.0 4.7 4.5 4.4 4.3 4.2 4.1 4.(
_ pH of Precipitation
1
• Figure 3-51. Nomograph to predict the pH of lakes given the sum of
^ nonmarine calcium and magnesium concentrations (or
nonmarine calcium concentration only) and the nonmarine
• sulphate concentrations in lake water (or the
• weighted-average hydrogen ion concentration in
precipitation) (Henriksen 1982).
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3-202
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Henriksen (1980) concluded that the nomograph could successfully
predict lake pH changes in response to changes in the pH of the
precipitation of the particular composition for that area and, if the m
titration process of lake acidification is reversible, the nomograph •
could be used to indicate the amount of decrease in precipitation
acidity necessary to restore acid lakes to bicarbonate lakes. _
A number of assumptions and cautions pertain to the use of the ™
predictor nomograph. One assumption initially inherent in the
predictor nomograph was that increases or decreases in the acidity of 4
precipitation do not affect the rate of leaching of Ca^+ or |
Mg2+ from soils. As Henriksen (1980) noted, this is a matter of
debate (e.g., see Aimer et al. 1978; Dillon et al. 1979) and a m
question that "certainly deserves further attention." If, for •
example, increased acidity of precipitation does cause increased
cation leaching from soils (instead of decreased lake pH), then the
titration hypothesis on which the nomograph is based is violated and •
extrapolations from the precipitation pH axis will be incorrect. ™
Henriksen (1982) has performed further research on this particular 4
problem. Using data from lakes in Norway, Sweden, Canada, and the 41
U.S., he: (1) compared historic and recent concentrations of
(Ca* +Mg*), and (2) evaluated ranges of (Ca* + Mg*) concentrations •
in lakes in similar geologic settings over a gradient of acidic •
deposition. In some cases he found that (Ca* + Mg*) concentrations
increased in conjunction with higher levels of acidic deposition. In
other cases he found no such concurrent increases. For the data he V
examined the maximum value of a "base cation increase factor" for the •
lake waters would be about 0.4 yeq (Ca* + Mg*)/yeq 804* (Henriksen
1982). Thus, estimates of the effect of changes in acidic deposition •
on the chemistry of lake waters still require knowledge of the degree |
of increase of base cation concentrations, ranging from 0 ueq
(Ca* + Mg*)/yeq 804* to roughly 0.4 yeq (Ca* + Mg*)/yeq 804*. •
This applied to certain lakes in Sweden, Norway, and North America •
where there was enough historical information to make an estimate.
However, he does state (p.38, Henriksen 1982) for Lake
Rishagerodvatten, Sweden, the factor was 0.63, and the Birkenes model •
(Christophersen et al. 1982) predicts an increase factor of about •
0.55. Dickson (1980) showed increases greater than 0.4 for some
Swedish west coast lakes. •
The increase factor represents possible responses of the watershed
system to acidic deposition. It reflects the geologic and hydrologic ^
sensitivity of the system. The lowest limit of the increase factor •
is zero, which refers to a system with little base exchange capacity *
in the organic soil, quartz (Si02) sands for the mineral soil,
and/or a lake in which precipitation does not flow through soils. •
Perfect seepage lakes without any drainage area other than lake area V
would qualify as systems with near-zero increase factors based on the
lack of flow through neutralizing soils. The maximum upper limit •
would be a watershed with calcareous soils or bedrock which would |
serve as a perfect buffer and yield an increase factor of 1.0 yeq
(Ca + Mg)/peq 8042~. _
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3-203
The leaching of aluminosilicate minerals in response to hydrogen ion
attack has been studied in the laboratory. Wollast (1967) found a
dissolution increase factor of 0.33 initially with respect to
hydrogen ion attack in 5% K-feldspar solutions. Furrer and Stumm
(1982) found a 0.4 increase factor in the dissolution of A^C^.
The factors that control a watershed's neutralizing capacity, and
hence the cation increase factor, are not well known and are
critical.
A second caution noted by Henriksen (1980) is that the predictor
nomograph should not be applied to waters containing high
concentrations of organic acids. Not only may the organic acids
affect lake pH in a manner independent of precipitation acidity, but
also ionic Ca^+ and Mg^+ may be overestimated inasmuch as analyses
for these ions include Ca^+ and Mg^+ bound to organics
(Henriksen 1980). A final point to note is that the derivation and
verification of this model is based upon the premise that the
observed data represent steady state conditions, both for
concentrations and pH in deposition, and concentrations and pH in
lake water.
A key question is whether the "predictor nomograph" is applicable to
sensitive lakes in northeastern North America. Relationships between
Ca* and (Ca* + Mg*) and between concentrations of these cations and
80^* may be different in regions of varying geochemistry in North
America. Furthermore, the empirical relationship between SO^* in
lake waters and 804* in precipitation (as well as the relationship
between SO^* in precipitation and pH of precipitation) may vary in
different geographical regions. Therefore, for more accurate
predictions it would be appropriate to develop region-specific
regression relationships and predictor nomographs like Henriksen's
from data bases for the regions of interest. Such studies would be a
useful extension of Henriksen's model and should be pursued.
Church and Galloway (1983) examined data from two small oligotrophic
headwater lakes in the Adirondacks and found, using only the (Ca* +
Mg*) and lake water (SO^*) axes, that the nomograph correctly
predicted the pH for all 66 measurements in a "bicarbonate lake" and
71% of 78 measurements for an "acid-transition lake". However, they
also found that the relationship between the precipitation pH axis
and lake water (864*) axis for the Adirondacks differs
significantly from the relationship for southern Norway. This is
possibly due to the different contributions of nitric and sulphuric
acids to precipitation acidity or to the presence of other cations in
precipitation in the two geographic regions. The variation of nitric
and sulphuric acid contribution to acidity of precipitation has been
further shown by Barrie (1982). Because as shown in Section 3.9.1,
nitrate has only minor influences on long-term acidity of the aquatic
regime in comparison with sulphate, only the relationships to
sulphate loading are considered in this section. For water pH values
less than 4.7, the presence of aluminum or of other buffering
apparently becomes important as shown by Henriksen (1980, 1982) and
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3-204
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may affect regression lines. However, we are more concerned with the
"transition" sector of the Henriksen nomograph.
Raines and Akielaszek (1982) examined data from 122 New England lakes •
in relation to the predictor nomograph. The nomograph correctly
predicted 6 of 7 lakes in the pH range <4.7 but incorrectly predicted flj
that 19 other lakes with higher pH values fell in this grouping. The £
nomograph correctly predicted 5 of 14 lakes that fell in the pH range
4.7 - 5.3 but incorrectly predicted that 32 lakes not in this range M
had such pH values. Of the 101 lakes in the pH range >5.3, the •
nomograph correctly predicted 60%.
For those New England lakes the nomograph predicted the pH of acidic •
lakes correctly but frequently predicted lower pH values than were •
actually observed in higher pH lakes. These differences may occur
because the relationships of calcium, magnesium, and sulphate are M
different in New England than they are in Norway, where the model was •
developed. Application of the predictor nomograph in New England
should be based on empirical relationships that exist in this region. —
Presently the relationship between lake sulphate concentration and m
atmospheric sulphate deposition has not been established for this ™
region.
Keeping in mind the important limitations and assumptions inherent in ||
its use, we have attempted an application of this approach to
estimating the effects of SO^" deposition on the chemistry of M
lakes in northeastern North America. Numerous lakes in Norway have •
calcium concentrations less than 50 yeq/L, and Bobe"e et al. (1982)
found that 7.5% (15) of 199 lakes sampled on the Precambrian Shield
of the Province of Quebec had calcium concentrations less than •
50 yeq/L. Raines and Akielaszek found that 11% (25) of 226 lakes and ™
streams in New England had calcium concentrations less than 50 yeq/L.
This indicates that such a limit would include all except the more M
sensitive waters. From the regressions given by Henriksen on: |
(1) the relationship of both (Ca* + Mg*) vs. alkalinity and (Ca*) vs.
alkalinity I and thus (Ca*) vs. (Ca* + Mg*)I (Henriksen 1980), and -mt
(2) the relationship of strong acidity to 804* and (Ca* + Mg*)I •
both with and without increased leaching of base cations (Henriksen
1982)1, we can roughly estimate a 804* concentration that yields a
pH of 5.3 in surface waters having initial Ca* concentrations of •
50 ueq/L. Using the regression given by Henriksen (1980) on the w
relationship of lake 804* concentration to 804* concentration in
precipitation and assuming an annual rainfall of 100 cm, we can then fl
estimate loading rates consistent with maintenance of a pH of 5.3 or ||
greater (pH 5.3 is the upper limit of Henriksen's transition zone).
The results of such calculations and the regression equations used •
are given in Table 3-32. Estimated loading values of wet sulphate •
deposition that will maintain lakewater pH at values 2:5.3 range from
approximately 26 kg/ha.yr (assuming no increased leaching of base
cations) to approximately 43 kg/ha.yr (assuming leaching of base •
cations of 0.4 times the change in excess sulphate concentration (see •
Henriksen 1980) and an initial lake 804* concentration of
0 yeq/L). •
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TABLE 3-32.
3-205
CALCULATION OF WET SULPHATE LOADINGS CONSISTENT WITH pH 5.3
OR GREATER IN LAKES WITH INITIAL CALCIUM CONCENTRATION OF
50 yeq/L OR GREATER (Regressions are from Henriksen [1980,
1982])
All Units (yeq/L)
(except where noted)
Ca*i
(Ca* + Mg*)i
(Ca* + Mg*)
(so4*)w p
(S04*)
(S04*)£ (kg/ha. yr)
No Leaching
of Base
Cations
50
70
70
81
53
26
Condition
Leaching
(according
.4
50
70
128
146
87
43
of Base Cations
to Eqn (4) below)
.2 .1
50 50
70 70
121 114
138 130
83 78
41 39
Ca*i
(Ca* + Mg*)±
(Ca* + Mg*)p
(so4*)w
(S04*)p
(so*)L
concentration of excess sulphate in lake
water prior to "acidification" (i.e.,
initial S04* concentration)
initial excess calcium concentration
initial sum of excess calcium plus excess
magnesium concentrations
predicted sum of excess calcium plus excess
magnesium concentration
final concentration of excess sulphate in
lake water
concentration of excess sulphate in
precipitation
areal wet sulphate loading assuming
annual rainfall of 100 cm
Equations Used in Calculations
= 1.32 (Ca*)± + 4.3 (adapted from
Henriksen 1980)
= [1.01 (Ca* + Mg*) + 1.81/0.9 (assuming no
leaching of base cations; Henriksen 1982)
= [1.01 (Ca* + Mg*)p + 1.8J/0.9 (assuming
maximal leaching of base cations; Henriksen
1982)
= (Ca* + Mg*)± + 0.4 ( S04*)w (Henriksen 1982)
= (Ca* + Mg*)i + 0.4 (S04*w -
(1) (Ca* + Mg*)±
(2) (S04*)w
(3) (S04*)w
(4) (Ca* + Mg*)
(5) (Ca* + Mg*)p
Substituting Equation 5 into Equation 3 and solving for (S04*)w yields
(6) (S04*)w
(7) (S04*)
(8) (S04*)£
= 2.04 (Ca* + Mg*)± + 3.64 - 0.82 (S04*)i
= KS04*) + 191/1.9 (Henriksen 1980)
= (S04*) /2 (assuming 100 cm annual rainfall)
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3.9.3.2 Cation Denudation Rate Model (CDR)
Thompson (1982) developed a model relating the pH of a river to the •
atmospheric loading of excess sulphate and the rate of cations from
a watershed via runoff (the Cation Denudation Rate or CDR). This
model is designed to apply to areas with acid-resistant bedrock, fl
till, and soils and relatively unbuffered surface waters. V
In most fresh waters the sum of base cations (Ca+2, Mg+2, Na+, K+) •
closely approximates the sum of anions HC03~ and SO^" after •
correction for seasalt or road salt contributions. Thompson (1982)
noted that if excess sulphate concentration is plotted against the ^
sum of the cation concentrations, a series of lines can be generated, •
each line representing constant bicarbonate concentration. If the ™
partial pressure of CC>2 (Pco2) ^n tne surface water in
question is constant, then each line also represents constant pH. •
This model may be applied to either rivers or lakes (Thompson 1982; fl
Thompson and Hutton 1982). If a runoff value of 1 m/yr is assumed
and the concentrations of terms in the axes of Figure 3-52 are ^
multiplied by this value, the axes become loading rates, and the •
figure becomes a plot of cation denudation rate (CDR, meq/m^.yr)
versus the discharge rate of excess sulphate (Thompson 1982). If all
the atmospheric sulphate deposited on the watershed is contained in •
runoff and if we assume that all non-seasalt sulphate comes from W
atmospheric loading, then the abscissa is equivalent to atmospheric
loading of acid sulphate. Note that if wind-blown dust has
neutralized some of the sulphuric acid in atmospheric deposition, the
loaing terms in Figure 3-52 must be corrected for these neutral
salts. Thus, according to the model, if CDR, runoff, excess sulphate _
load, and Pc02 are known, mean pH can be estimated. •
An example of how model calculations are made is given below. If the
rate of excess SO^" loading is less than the CDR by 20 meq/m^.yr •
(i.e., the HCC>3~ residual equals 20 meq/m2.yr), the model estimates •
that the resultant runoff water (assuming a yield of 1 m/yr) will
have a mean pH of 5.6 (Figure 3-52). As the rate of excess •
504^" loading approaches the CDR, the runoff water will approach a f
pH of 5.1 (at which HC03~ alkalinity is totally exhausted). Data
for very soft water rivers in Nova Scotia and Newfoundland that have _
mean runoff rates near 1 m/yr are shown in Figure 3-52. These rivers •
have a total CDR ranging from 55 to 200 meq/m^.yr. In 1973 at *
least three of these rivers received SO^" loads exceeding
their CDR and had median pH values less than 5.1. M
At first glance the CDR model appears to be quite similar to the
predictor nomograph of Henriksen. The CDR model is developed •
strictly from consideration of charge balance, however, whereas the •
predictor nomograph is strongly dependent on empirical observations.
Thompson (1982) explicitly assumes that CDR is independent of acid —
loading; that it varies only with discharge. The recent data review •
by Henriksen (1982) shows that CDR cannot be considered to be ™
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3-207
CO
r2.5
PC02=10
RUNOFF = 1m/yr
100
200
ACID LOAD or EXCESS SO4
(meq/m2. yr) (|aeq/L)
2-
Figure 3-52. The model plot - pH predicted for consideration of the
sum of cations and sulphate (modified from Thompson
1982).
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3-208
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constant in all cases. Thompson et al. (1980) compared data from
between 1954-55 and 1973 for very soft water rivers in southern Nova
Scotia. They found a lower pH and higher excess 804^" •
concentrations in the most recent data but did not find significant |
changes in major cation loads.
A way in which the CDR model is similar to Henriksen's predictor •
nomograph is that it does not apply in situations where organic acids
strongly influence pH. The CDR differs, however, in that it does not
consider the possible effects of buffers other than bicarbonate. •
Also, PCOZ roust be known to estimate pH with the CDR model.
As is commonly known, Pcc-2 varies significantly in surface
waters. A
Raines and Akielaszek (1982) applied the CDR model to data from 122
New England lakes. The CDR model gave better results than the •
predictor nomograph (discussed above). Predicted pH agreed very well •
with measured pH at values <_6.3. However, this model also predicted
lower pH than was measured for many lakes with pH >6.3.
As discussed above, estimates of the relationship between sulphate H
deposition rates and surface water pH may be made. As an example,
the Roseway River, Nova Scotia (Figure 3-53) has a CDR of •
56 meq/m^.yr. If all of the assumptions noted above hold and if •
the acidification process is reversible, then a reduction of the
sulphate loading rate to 40 meq/m^.yr (20 kg SC>42-/ha.yr)
might be expected to return the river to an annual pH of roughly 5.3. •
A significant problem exists with such a prediction. The Roseway
River has strongly coloured waters, as do the Mersey and Medway
Rivers (also shown in Figure 3-53). As Thompson (1982) notes, the pH II
values of these rivers "have been thought to be dominated by •
naturally- occurring organic acids." Thompson (1982) feels that
"their low pHs can be explained quite well on the basis of simple •
inorganic chemistry." No chemical data (e.g., Gran titrations for •
weak and strong acids) were presented to confirm this. If the pH
values of these rivers were controlled by naturally-occurring organic _
acids, reduction of excess sulphate deposition would not result in •
the increases in stream water pH predicted. ™
Figure 3-54 and Table 3-33 were calculated based on the Thompson •
(1982) model. If 80 yeq/L of cation concentration (roughly f
equivalent to 50 ueq/L Ca2+ as used in Table 3-33) is used as a
criteria for basin sensitivity to acidification, maintenance of the m
basin water to a mean pH >5.3 should be possible with sulphate •
loadings of 35 kg S042~/ha.yr given a runoff of 100 cm/yr.
Other combinations of sulphate deposition and runoff are shown on
Table 3-33. It should be noted that any retention of sulphate within •
the watersheds or increased leaching of base cations would violate 9
assumptions in the model, causing the above loading estimates to be
too low. •
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3-209
200
CT
0)
§
rr
Q
<•> 150
100
RIVER CDR EXCESS SO42~ MEDIAN pH
Wallace 203
Meteghan 129
Le Have 126
Pipers Mote 101
St. Mary's 100
Tusket 75
N£.Pond 71
Medway 71
Mersey 66
Roseway 56
O 4.3
Figure 3-53.
50 100 150
Excess SO4 meq/m2- yr
Cation Denudation Rate Model applied to rivers of Nova
Scotia and Newfoundland (Thompson 1982).
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3-210
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2 400
CO
CO
£ 360
3
0
C
C
o
**
CO
l_
•frrf
c
0)
o
c
o
o
c
o
4~
CO
o
to maintain Aquatic Regime at:
__ pH 5.3, i.e., HCOg = 10jjeq/l_
pH 5.8, i.e., HCO = 32jueq/|_
320
Runoff
30 cm/yr
TJ
0
o 280
CD
i_
i_
O
3 240
£ 200
3
CO
o 160
«
•S 120 -
80
40
0
8
12
16 20
24 28
32
36
40 44
Excess Sulphate (kg SCL2 /ha - yr)
Figure 3-54.
Relation of excess sulphate and cation concentration
for pH 5.3 and 5.8 for basin runoff of 30, 50 and
100 cm/yr. The model was developed for an area with
100 cm runoff. It has not been corroborated for areas
with lower runoff (derived by the Work Group from
Thompson 1982).
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3-211
TABLE 3-33. ACIDIFICATION SENSITIVITY OF SURFACE WATERS RELATED TO
SULPHATE LOADING FOR TWO pH OBJECTIVES AND THREE RUNOFF
Thompson (1982)]
Cation
Concentration
( jjeq/L) pH Objective
300 5.3
5.8
200 5.3
5.8
100 5.3
5.8
50 5.3
5.8
Runoff (cm/yr)
30
44
40
28
25
13
10
6
3
50
50
50
47
42
22
17
10
5
100
50
50
50
50
45
34
20
9
• VALUES [derived by the Working Group from CDR model,
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The model was developed for an area with 100 cm runoff. It has not
been corroborated for areas with lower runoff.
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3-212
3.9.3.3 Summary
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The application of two simplified models to the problem of relating •
wet deposition of sulphate to lake pH has been discussed in this •
section. Before any environmental or water quality model can be used
to make estimates with specified confidence of future conditions in a •
particular geographic region, the applicability of that model for •
that region and conditions must be verified. This process of
verification is just beginning for Henriksen's predictor nomograph M
and CDR model to northeastern North America. Until such verification •
(and perhaps, model adaptation) is achieved, quantitative predictions
based on these models must be viewed with caution.
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3.9.4 Summary of Empirical Observation and Modelling
to normal or altered fluxes in the hydrologic regime; the regional
responses in lake chemistry; the basin characteristic which influence ^
sensitivity to acidification; evidence of changes or trends in •
surface water quality in sensitive regions; evidences of alteration ™
of biological components; and finally, the methodologies which are
available to assist in estimation of target loadings by atmospheric B
deposition which would be consistent with protection of the ecosystem •
to a degree acceptable to society. Because environmental concerns
are of rather recent recognition and those which have been recognized M
are most often related to more intense urban contamination, long term •
records of verified significance are available in only a few cases
from which firm conclusions can be drawn relating to acidification of
remote ecosystems. A deterministic knowledge of the inter- •
relationships of the bio-hydrogeochemical system and of its responses ™
to altered precipitation chemistry is not yet available, therefore
rendering precise predictive modelling of system responses, as yet, fl|
unattainable. These limitations have been thoroughly reviewed in I
recent summaries of the Associate Committee on Scientific Criteria
for Environmental Quality, National Research Council of Canada m
(Harvey et al. 1981) and by the Committee on the Atmosphere and the I
Biosphere, Board on Agriculture and Renewable Resources Commission on
Natural Resources (NAS 1981) and are further detailed in this report.
However, these learned summaries of present knowledge have all •
indicated strong evidence of significant ecosystem deterioration due 9
to past and present levels of acid precipitation loading and thus
indicate the urgent need to use this present knowledge to arrive at A
best estimates of levels of acid loadings which can be tolerated. |
While this chapter has considered only the aquatic portions of the «
ecosystem, it would appear that because of the interactions with •
other components, protection of the aquatic regime would, to a
large degree, result in protection of the total environment. This
sub-section therefore, reviewed the information and methodologies •
presented earlier with respect to their utility in producing •
estimates of loading/response relationships.
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3-213
As developed in previous sections, acidification of aquatic regimes
can be related to proton (H+) loading, concentration of IT1", i.e.,
of precipitation, or to the constituents of the loading which
determine the acidity (i.e., the major ionic species). The
anthropogenic loadings add to and interact with the natural
components to an extent that also influences the factors available
for effective control. Evans et al. (1981), after reviewing the
extensive evidence of dose response acidification relationships and
considering the empirical model approach of Henriksen (1980) have
proposed that "an annual volume weighted H+ concentration of 25
yeq/L (pH 4.6) in precipitation appears to be a critical threshold."
These authors have reached their conclusions through the basic
consideration of H+ exchange in the reaction processes and through
general empirical observations of dose response in sensitive regimes.
However, as reviewed in earlier sections, the biosystem response and
ability to assimilate nitrate, ammonia or sulphate (the primary
acidifying ions of precipitation) differ and therefore the acidifying
potentials of these ions differ. In addition, Stensland (1979) and
Barrie (1981) have shown that the ionic concentrations of
precipitation over eastern North America varies as to relative
contribution to its acidity in both space and time. Thus the H+
concentration cannot be considered to have a unique relation to the
acidity controlling ions nor has it a unique relation in its dose
response in the bio-hydrogeosystem. Thus, neither H+ concentration
of precipitation nor H+ loading rates form acceptable criteria for
target loadings in relation to protection of aquatic ecosystems from
acidification.
Henricksen (1980) has argued that surface water acidification can be
accounted for as the titration of bicarbonate waters and replacement
of bicarbonate by sulphate in the ionic charge balance. He found
good empirical agreement between sulphate loadings and observed
acidification in widely diverse areas without consideration of any
nitrogen species. His relationship to precipitation pH, as cited by
Evans et al. (1981), was empirical and based upon Norwegian
precipitation and was not an integral part of the argument. It
should be stressed here, that while Henricksen's model has a basis in
chemical equilibrium, as shown by Thompson (1982), it is in fact a
"phenomenological" model which derives from actual dose response
observations.
A range of sulphate loading vs bio-geo-system responses observed in
eastern North America are summarized in Table 3-26 and Summary Table
(p. 3-178). This includes several cases relating to episodic event
pH changes. While the number of cases are small and statistical
significance cannot be assigned, the identified cases of surface
water acidification and observed biosystem effects all fall within
regions of sulphate deposition of greater than 17 kg S042-/ha.yr.
There appear to have been no reported cases of identified
acidification which cannot be related to organic sources in areas of
less than this level of sulphate deposition.
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3-214
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The Canadian members of the Work Group consider that this evidence,
often circumstantial but not inconsistent with theory, leads to the
approach best able to provide estimates of target loadings of •
sulphate in relation to surface water acidification. It is well gj
recognized that this estimate is of limited accuracy in terms of
predicted ecosystem response and must surely be subject to later ^
re-evaluation as more information is developed from scientific study. •
The empirical observations presented in Table 3-26 and Summary *
Table (p. 3-178) immediately suggest a target loading of sulphate
which could be accepted but is only poorly defined in terms of •
geosystem parameters. The Henricksen-Thompson model permits a I
quantification of the target loadings in terms of the geochemical
r)~i- Q I
Jbasin sensitivity parameter CDR (Ca* + Mg^ ) as an approximation or •
unaltered alkalinity as suggested in Section 3.9.3 and further •
developed in this section. As pointed out by Henricksen (1980) this
model will have, perhaps, significant errors below the titration end —
point for alkalinity due to other buffers but should apply with •
sufficient accuracy for estimates of the loadings of sulphate which ™
would control the aquatic regime acidity above this transition pH.
The CDR serves as the basic geosystem sensitivity criteria in this •
model and thereby links the basin hydrology and the sulphate loading fl
to sensitivity to acidification. CDR and cation concentrations are
related through the hydrological runoff. M
Information derived from the Thompson (1982) model may, within
the limitations cited, be used to estimate target loadings of ^
sulphate (Figure 3-54 and Table 3-33). Thus if 200 yeg/L of cation •
concentration (also unaltered—at/cai-inity; is used as a criteria for ™
basin sensitivity to acidification, protection of the basin water to
a mean pH of 5.3 would be indicated for sulphate loadings fl
47.5 kg S0^2~/ha.yr if runoff of 50 cm/yr occurred. For a m
30 cm/yr runoff the protection would only tolerate a loading of
28 kg SO42~/ha.yr. Thus the criteria of 200 yeg/L total M
cations or unaltered alkalinity is a reasonable choice of threshold •
of sensitivity to acidification over much of eastern North America
where runoff may be near 50 cm and sulphate loadings exceed
40 kg S042~/ha.yr (see Figure 2-6b). •
A target loading of 15-20 kg SO^2~/ha.yr would, by this model,
serve to maintain surface water pH greater than 5.3 on an annual H
basis for basins having cation concentrations of 200 yeg/L or greater ^
even in areas of low runoff. More sensitive basins in low runoff
areas could not tolerate this level of loading and maintain a pH m
greater than 5.3. •
The estimates of dose-response relationships presented here do not
account for the episodic events discussed earlier which may, in some •
ecosystems, be cause for more concern than that based on the mean ™
acidity. The estimates do not consider any time response and must
therefore be limited to steady state conditions. Rate of response of •
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3-215
any basin to changes in precipitation loading, either quantity or
quality, must relate in general to the water residence time. Other
factors such as ionic migrations in the soils are not considered.
Thus no response rates or equilibrium times are implied, in any
sense, by these loading estimates.
In the watershed studies summarized above, sulphate in precipitation
was used as a surrogate for total acid loading. Sulphate in
precipitation is reliably measured. It is recognized that dry
deposition of sulphate and sulphur dioxide, and the wet and dry
deposition of nitrogen oxides, nitric acid, particulate nitrate and
ammonia, as well as other compounds also contribute to acidic
deposition. Based on documented effects, wet and dry deposition of
sulphur compounds dominate in long-term acidification.
Based on the results of the empirical studies, interpretation of
long-term water quality data, studies of sediment cores and models
that have been reviewed, we conclude that acidic deposition has
caused long-term and short-term acidification of sensitive surface
waters in Canada and the U.S. The work group also believes on the
basis of our understanding of the acidification process that
reductions from present levels of total sulphur deposition in some
areas would reduce further damage to sensitive surface waters and
would lead to eventual recovery of those waters that have already
been altered chemically or biologically (Loss of genetic stock would
not be reversible.)
The U.S. members conclude on the basis of modelling and empirical
studies that reductions in pH, loss of alkalinity, and associated
biological changes have occurred in areas receiving acidic
deposition, but cause and effects relationships have often not been
clearly established. The relative contributions of acidic inputs
from the atmosphere, land use changes, and natural terrestrial
processes are not known. The key terrestrial processes which provide
acidity to the aquatic systems and/or ameliorate atmospheric acidic
inputs are neither known nor quantified. The key chemical and
biological processes which interact in aquatic ecosystems to
determine the chemical environment are not known or quantified.
Based on this status of the scientific knowledge, the U.S. Working
Group concludes that it is not now possible to derive quantitative
load ing/effects relationships.
3.10 CRITICAL RESEARCH TOPICS
The following topic areas represent issues in which there are major
information gaps, and which should be addressed by research programs,
in both the U.S. and Canada, at the earliest possible date.
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3.10.1 Element Fluxes and Geochemical Alterations of Watersheds
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Three areas of research are needed here, all requiring relatively •
intensive study of both terrestrial (geochemical) and aquatic B
(hydrologic) components, mostly focused around calibrated watersheds
of comparable research design and intensive data quality assurance. —
•
B
1. The four ions of primary concern regarding acidification are
hydrogen, ammonium, sulphate, and nitrate. Each ion reacts
differently with the soil matrix and vegetation. It is
necessary, therefore, to define , in specific terms , the fate and
effect on surface water acidification of hydrogen, ammonium,
sulphate and nitrate ions originating as atmospheric input. m
Comparison of results from calibrated watersheds with different
soil and vegetation conditions is urgently needed. This report
indicates that priority may have to be given to sulphur B
emissions control, drawing heavily on evidence that nitrogen 9
deposition does not contribute significantly to long-term
surface water acidification, even though it contributes to
precipitation acidity and pH depression during snowmelt or
runoff events. The long-term necessity for a sulphur control
priority needs to be established beyond doubt, as soon as M
possible, in order to minimize the risk of making costly errors B
in a control program.
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2. Acidic deposition results in mobilization of metals, such as B
aluminum, iron, zinc and manganese, from the soil particles in B
watersheds. Further work is needed to define the amounts and
species of metals leached from watersheds and their biological
consequences.
3. There is evidence that groundwater is being acidified, and that M
metal concentrations are elevated, in areas where snowmelt gains B
direct access to sandy subsoils with low acid neutralizing
capacity. The effect may be seasonal, with pH values recovering
during the summer, as neutralization slowly takes place. B
Further surveys are needed to establish the extent and B
characteristics of groundwater modification over time and across
geographical gradients in acid loadings. •
3.10.2 Alterations of Surface Water Quality •
Two major areas of information needs have been identified in the
extent and periodicity of surface water quality effects:
1. The geographical extent of surface water acidification is not B
yet fully documented in North America. Obvious data gaps exist
in the central, southern and western U.S. and in parts of •
Canada. In addition, reliable data on time-trends in water £
quality appear to be sparse throughout North America, although
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3-217
some data have not yet been evaluated. Much of the new data
needed can be obtained as part of the long-term monitoring
program described below. The critical need is to begin long-
term water quality measurements, in a carefully selected range
of aquatic environments, as soon as possible.
2. One of the most common manifestations of acidic deposition
observed in eastern North America is periodic pH depression in
streams and lakes, due to snowmelt or heavy rain. Since
periodic low pH is a current problem for biological resources,
and likely to remain so until acid deposition is reduced, the
quantitative relationship between acid deposition and short-
period pH depression should be determined for a broad spectrum
of aquatic environments. A dose-response relationship for
episodic acute exposures to H+ and aluminum will be a major
element in defining acceptable acid loadings.
3.10.3 Alteration of Biotic Components
Effects on the biological components of aquatic ecosystems are known
only partially. Five research topics are identified:
1. It is essential that the biological responses to various water
chemistry changes induced by acidic deposition, be evaluated in
considerable detail to define dose-response relationships
further. Studies of dose-response relationships in aquatic
ecosystems should include surveys of phytoplankton, macro-
phytes, zooplankton, benthos and amphibians. Several species
among these groups are quite sensitive to changes in pH.
Of particular importance to the dose-response relationship is
quantification of response data from indigenous species which
may be vulnerable to low pH or elevated aluminum, and the pH at
which effects are expressed. Special attention needs to be
given to determining the pH at which species unique to certain
areas are harmed and begin to show some failure in reproduction.
In addition, community-level attributes of aquatic systems are
likely to be sensitive to acid-induced stresses, but are
difficult to determine; nevertheless, they should be understood
fully. These include plankton species composition, predator-
prey relationships, and trophic-state modification of lakes due
to altered nutrient cycles.
2. Damage to fish populations is of particular concern because the
loss of fish breaks a major link of the water/terrestrial food
chain. Sport fishing is an important industry in most of the
areas affected by acidic precipitation and reduction in fish
supply could have serious economic consequences. Mechanisms by
which low pH and high metal concentrations affect fish should be
studied to improve general understanding of the toxicity
phenomenon and to improve the ability to predict future effects
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3-218
and if so, whether there has been any reduction in spawning
success for fish species in those tributaries.
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if acidic deposition continues. Fish sensitivity to H+ and
metal ions should be determined, by direct bioassay, at
different stages in the life cycle, concentrating on fl
reproduction and recruitment. Behavioural or physiological •
changes (e.g., blood ion levels) known to be affected by
sublethal acid and metal concentrations should also be •
evaluated. Long-term monitoring should include fish population •
data, as well as other measures of biological productivity.
3. Further study is needed to define the biological effects and •
tolerances for periodic pH depression in streams and lakes. •
Current work should be extended , to include the Great Lakes
tributaries draining Precambrian areas. All such potentially B
sensitive areas in the U.S. and Canada should be surveyed, to |
determine whether low pH and high metal concentrations occur,
•
•
Mercury concentrations in fish and other wildlife may be
increased by the acidification process and/or by increased
atmospheric emissions. Increased effort should be placed on
measuring existing mercury concentrations and time trends
throughout the wildlife food chain, as a function of lake and
stream pH values. Laboratory and field studies are needed to
establish the biological significance of various mercury
concentrations in indigenous species of fish, birds and •
mammals. I
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5. When aquatic and/or terrestrial productivity is affected, the
effect is often evidenced through the entire food chain. Thus, •
there is reason to believe that acidification will have an •
adverse effect upon food availability to the higher trophic
levels of the food chain, including aquatic birdlife and •
mammals. The long-term effects of habitat damage on the |
populations of wildfowl and other wildlife should be better
defined, and the losses of habitat should be quantified. •
3.10.4 Irreversible Impacts
1. Geochemical and hydrologic principles suggest that the processes W
of sulphate accumulations, and associated acidification of soils
and surface waters, represent a large-scale titration of •
available acid neutralizing capacity. There is evidence that f|
the capacity of watersheds to provide neutralization of acids
may become depleted, over long periods. Therefore, further work _
is needed to define the rate of acidification of surface waters, •
develop predictive models to quantify watershed capacity to
neutralize acid over the long term, and to anticipate recovery
following abatement. •
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The studies should include measurements on the rates of
acidification of lake and stream sediments. The results of such
studies are needed to assist in setting acid loading tolerances
which will be protective of the aquatic environment in the long
term.
3.10.5 Target Loadings and Model Validation
Much uncertainty remains as to the quantification of sulphate
deposition level ("target loadings") consistent with no further
significant degradation of natural resources. Two areas of research
are needed:
1. Several relationships, based on field environmental data, have
been used to develop descriptive and predictive models of the
acidification process. Dickson's relationship, the Henriksen
nomograph, and the episodic receptor/dose relation, appear to be
potentially useful empirical models which warrant comparative
analysis with similar background data bases. Efforts should be
made to conduct additional validation of existing and emerging
model descriptions of the process of acidification.
2. Relatively detailed simulation models of the acidification
process, and its effects, are being developed by several
research groups. These should be evaluated, using watershed
data bases from a number of intensive study sites in sensitive
areas, as identified in this report. If important data are
presently missing at these sites, they should be added to the
measurement program, or if certain summaries are not being made,
these should be added. The need is to have the most complete,
quantitative long-term dose-response models evaluated fully and
compared with the more empirical field relationships now in use.
In support of this validation process, every effort should be
made to maximize the use of existing information from all
sources.
Reasonable validation of both types of models will require
considerable new research. Study areas for evaluating atmospheric
transport models (see Work Group II report) and loading predictors
should coincide with detailed studies of sensitive receptor areas.
Locations which already have some data, and which should be
considered, include:
Experimental Lakes Area - Ontario
Boundary Waters Canoe Area Wilderness - Minnesota
Algoma Area Watershed Study - Ontario
Dorset-Haliburton Study Area - Ontario
ILWAS Project - New York
Laurentide Park (Lac Laflamme) - Quebec
Kejimkujik Park - Nova Scotia
Hubbard Brook - New Hampshire
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Northern Highlands Lakes - Wisconsin
Coweeta - North Carolina
Andrews - Washington
North Cascades - Washington
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3.10.5.1 Long-Term Data Collection and Monitoring •
The present limited ability of the scientific community to assess
critically the extent of impacts from elevated acidity in H
precipitation, and from other components of atmospheric deposition, |
is a consequence of few reliable baseline observations on sensitive
aquatic environments. This lack of systematic data arises, •
primarily, because many studies and monitoring programs were planned •
to define the influences of local anthropogenic development and are,
therefore located near these influences. Because acidification is of
greatest importance in remote areas unaffected by local discharges, •
very few areas exist with any long-term baseline information. ™
Filling this information gap as quickly as possible should be a B
priority in both the U.S. and Canada. This information is needed so |
that positive, definitive analyses of ecosystem response to the
changes in atmospheric deposition can be carried out, with extensive M
verifications. Unless a monitoring program is in place and providing •
a documented time-series of system properties, there will be no
significant capacity to quantify the results of either emission
reductions or increases. •
While a variety of data needs have been implicit throughout the
aquatic effects section, certain classes of long-term measurements
are needed at selected sites. Included are the following four:
1. Since a major component of aquatic research is the calibrated «
watershed, long-term studies of these systems should be •
intensified with the general objective of improving the
estimates of rates of changes in water quality and biological
effects relative to acid loadings (i.e., dose-response •
relationships), improving the understanding of the relative ••
influence of sulphur and nitrogen loading; and establishing
better measures of lake sensitivity, so that the present and
potential extent of the problem can be more clearly defined.
2. Analyses should be undertaken of all available baseline studies, •
including regional monitoring of surface water quality, •
plankton, fauna, soil, and vegetation records.
3. Criteria for selection of streams and lakes for new monitoring •
of water quality and biota should include factors related to •
alkanity sources, lake morphometry, watershed morphometry,
groundwater inputs, vegetation cover (i.e., age of forest and •
community structure), surface water chemistry, groundwater £
chemistry, and type of biotic community (cold water, warm water
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etc.). The regions and lakes chosen for analysis should range
from very sensitive, through moderately sensitive, to "tolerant"
(reference lakes), although a geographic grid of comparable
sites should also be developed. Data collected should include
chemical and biological parameters identified as susceptible to
change.
4. Experimental manipulations should be carried out, using adjacent
watersheds with small lakes. Watershed-level experiments should
include "simulated acid precipitation" additions of ff1",
SO^-, NH^, N03~, etc., so that long-term recovery, following
termination of acid additions, can be investigated.
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1974. Effects of acidification on Swedish lakes. Ambio
3:30-36.
1982. Effects on fish of metals associated with
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3.11 REFERENCES
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SECTION 4
TERRESTRIAL IMPACTS
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4-1
SECTION 4
TERRESTRIAL IMPACTS
4.1 INTRODUCTION
A number of air pollutants generated by various sources cross
international, state and provincial boundaries. The main pollutants
which are potentially harmful to terrestrial ecosystems are oxides of
sulphur (SOX), oxides of nitrogen (NOX), particulates, and secondary
products, such as oxidants and acidic deposition. There are also
smaller amounts of heavy metals, several of which have potentially
toxic significance after accumulation.
Sulphur dioxide (802) is emitted at phytotoxic concentrations by a
large number of mainly anthropogenic sources, including power plants
and smelters. Most of this S02 is deposited in dry forms near the
sources, though some is transformed chemically in the atmosphere to
other sulphur compounds. A moderate amount of S02 remains widely
distributed in the atmosphere. In areas remote from sources, the
concentration of S02 near the ground is close to background levels,
and not likely to cause adverse direct effects. However, S02 is
transformed in the atmosphere through a series of reactions into
sulphuric acid (H2S04) thus contributing to the formation of the
secondary pollutant, acidic deposition. Similarly, NOX gives rise
to nitric acid (HN03) and are likewise precursors of acidic
deposition. Ozone (03) is also an indirectly emitted secondary
pollutant formed in the atmosphere in the presence of sunlight, after
chemical transformations of nitrogen dioxide and reactive
hydrocarbons.
In summary, acidic deposition and ozone, although secondary in
nature, are usually considered to be long-range transported pollut-
ants as they frequently occur in relatively high concentrations at
distances hundreds of kilometres from the source of their primary
precursors. Because ozone is a strong oxidizer, oxidative decay
usually is rapid in polluted atmospheres and therefore decreases in
concentration during late afternoon and evening as sunlight intensity
decreases. However, ozone can persist overnight in rural areas or at
altitudes where there are low concentrations of reactive components
(Jacobson in press).
Improved understanding is needed of the ecological effects of the
phytotoxic primary and secondary pollutants on terrestrial eco-
systems. Field observations and laboratory studies have provided
detailed descriptions of the visible injury symptom syndrome produced
by ozone. Several review articles and chapters have provided
excellent descriptions of these symptoms (Brandt and Heck 1968;
Hill et al. 1970; USEPA 1978a). Field studies including the use of
field chambers (Heagle et al. 1973; Thompson and Taylor 1966) and
those with plots located in a natural ozone gradient have
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demonstrated that chronic ozone exposures suppress growth and reduce
yield, often in the presence of little or no visible injury symptoms.
A more detailed description of the response of plants to acute and •
chronic exposures to ozone is presented elsewhere (NAS 1977). •
It has been more difficult to determine the adverse or beneficial •
effects of acidic deposition on plant communities. Although |
simulated rainfall experiments have produced some direct effects on
plants exposed to higher than normal hydrogen ion (H+) loadings, ^
direct effects have not been documented conclusively in the field for •
vegetation exposed to ambient precipitation (Jacobson 1980).
However, some studies have demonstrated the direct effects of acidic
deposition on soils (Cronan et al. 1978; Dickson 1978). A
Indirect effects of acidic deposition (i.e., acting through soil,
other organisms) and its implications are even less well known. •
Increases in acidic deposition could result in accelerated changes in •
the natural evolution of soils, leading to alterations in soil
fertility over the long term. These changes in soil chemistry could ^
have detrimental implications for long-term sustained forest I
productivity, and also must be considered in association with aquatic ™
sensitivity.
This section on terrestrial effects of transboundary air pollutants 0
is presented in four parts: (1) effects on vegetation; (2) effects on
wildlife; (3) effects on soil; and (4) sensitivity assessment. Where M
possible, the information on acidic deposition and combinations of •
these pollutants has been partitioned and further subdivided into
agricultural crop and forest effects.
4.2 EFFECTS ON VEGETATION
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4.2.1 Sulphur Dioxide (S02)
4.2.1.1 Introduction
Sulphur dioxide is an air pollutant of concern to vegetation having
most often been recognized for inducing direct foliar effects to
plants growing proximal to major point sources of emission. The •
phytotoxicity of this gas has been studied extensively around V
long-term sources such as Sudbury, Ontario (Dreisinger and McGovern
1970; Linzon 1971) and the districts of Fox Creek and West •
Whitecourt, Alberta, (Legge et al. 1976). Controlled long-term •
exposure studies have recently been completed as part of the Montana
Grasslands Studies (Lee et al. 1978; Preston 1979). This pollutant _
has also been considered of great importance to the vegetation within •
the heavily industrialized areas of Great Britain (Cowling and Koziol *
1978) and central Europe (Guderian 1977).
Sulphur dioxide is not found on a regional basis at concentrations ^
sufficient to cause direct injury to most plant species. Long-term,
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4-3
low-dose studies have demonstrated direct effects to lichen
communities (Hawksworth 1971) and indirect effects to several plant
species (Keller 1978, 1980; Laurence 1978). Likewise effects may
result from lower doses of pollutants in combination with special
reference to 03 and S0£ in mixtures (Heagle and Johnston 1979;
Reinert and Nelson 1980). Several reviews of the effects of SC>2 on
vegetation are available (Guderian 1977; Jacobson and Hill 1970;
Linzon 1978; Rennie and Halstead 1977; Treshow 1970; USEPA 1973,
1978b).
4.2.1.2 Regional Doses of S02
As presented in Table 2-3 of Section 2 (Rasmussen et al. 1975),
estimates of global background concentrations of SC>2 in gaseous
form should be expected within a range of approximately
0.5-5.0 yg/m3 (0.0002-0.002 ppm S02 at STP) with expected
residency times of these concentrations to last from one to five
days. Regional S02 emissions are shown in Figure 4-1.
Mueller et al. (1980) reported on atmospheric pollutant data
collected during the period August 1977 - October 1978 for an area
covering much of the eastern half of the United States (Figure 4-2).
Monthly 1-hr averages varied from 5-40 yg/m3 (0.002-0.015 ppm
802). The highest annual average SC>2 concentrations occurred
along the Ohio River Valley; averages ranged from 0.019-0.029 ppm
S(>2. The maximal 1-hr concentrations were from 0.11-0.19 ppm S02
and occurred in the same area during October 1978. Hourly deposition
values of 1.5-2.3 ppm S02 are common near large emission sources
(USEPA 1978a).
In the northeast alone, anthropogenic sources exceed all others by a
factor of 12.5. Within this region, S02 levels annually average
16 yg/m3 (0.006 ppm S02) (Shinn and Lynn 1979) which is several
times that recorded in pristine areas. Therefore, it is reasonable
to assume that at the present time concentrations of S02 seldom
reach direct foliar injury thresholds for vegetation growing in
forested areas or in areas of significant agricultural production.
Duchelle and Skelly (1981) reported S02 concentration ranges of
0.001-0.002 ppm/hr S02 during the summer seasons of 1979 and 1980
within the Shenandoah National Park in Virginia and did not consider
this pollutant of importance to vegetation in the area.
Distribution of even these low doses of SC>2 (and N02) over the
major portion of eastern United States corresponds well with known
ozone occurrences (USEPA 1978b).
4.2.1.3 S02 Effects to Agricultural Crops
There are several possible responses to S02 and related sulphur
compounds: (1) fertilizer effects appearing as increased growth and
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4-4
I
>10,000
1000.1-10,000
x3 100.1-1000.0
&x
:-:+\ 10-100.0
(ANNUAL EMISSIONS IN g/s)
200 0 200 400
0 1
2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39
Figure 4-1. Magnitude and distribution of sulphur dioxide
(802) emissions in eastern North America. Data from
SURE II data base and Environment Canada (Environment
Canada 1981d).
1
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4-5
SO2(ppm)
Aug. 1977
Jan./Feb. 1978
SO2 (ppm)
Jul. 1978
SO2 (ppm)
Oct. 1977
1978
SO2 (ppm)
Oct. 1978
Figure 4-2. Geographic distribution of monthly arithmetic means for
S02 (Mueller et al. 1980).
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4-6
acids and proteins. The rate of entry is particularly important to
determining toxicity. Plants have an inherent, and apparently
I
1
yields; (2) no detectable responses; (3) injury manifested as growth
and yield reductions without visible symptom expressions on the
foliage or with very mild foliar symptoms that would be difficult to •
perceive as air pollution incited without the presence of a control •
set of plants grown in pollution-free conditions; (4) injury
exhibited as chronic or acute symptoms on foliage with or without •
associated reductions in growth and yield; and (5) death of plants £
and plant communities.
Sulphur.dioxide passively enters plants via stomata as part of normal •
gas exchange during photosynthesis. Many factors govern stomatal •
opening and closing including light, relative humidity, CC>2
concentration and water stress. Sulphur dioxide uptake and ingress ft
may also be limited according to plant genetics, previous exposure to V)
SC>2 (Jensen and Kozlowski 1975) and subsequent biochemical and/or
physiological alterations within exposed plants. Sulphur dioxide has •
been shown to increase or decrease stomatal resistance and this may •
directly affect potential for the photosynthetic performance
(Hallgren 1978). Based on the available literature, it is difficult
to assess the relationship of SC>2-induced biochemical and/or I
physiological changes at the cellular level in relation to subsequent •
effects on photosynthetic activity or resultant growth and yield.
Sulphur dioxide, upon absorption is further oxidized to 863 and
50^2- ancj subsequently is incorporated into S-containing amino
» *- A- A
species dependent, capacity to absorb, detoxify, and metabolically ™
incorporate SC>2 and some plants may absorb low concentrations of
S02 over long time periods without injury. •
Atmospheric S02 can have beneficial effects to agronomic vegetation
(Noggle and Jones 1979). Sulphur is one of the elements required for •
plant growth and Coleman (1966) reported that crop deficiencies of S •
have been occurring with increasing frequency throughout the world.
Several studies using SC>2 as a nutrient supply for S requirements •
of plants have been accomplished under varying degrees of soil- •
sulphur availability (Cowling et al. 1973; Faller 1970; Noggle and *
Jones 1979). The results of these and other studies leave little
doubt that application of S as a nutrient via SC>2 fumigation of •
plants grown on borderline or S-deficient soils will lead to V
increased productivity.
The interpretation of studies demonstrating such beneficial effects •
must be evaluated in light of their single influence to one crop.
Long-term natural ecosystem studies showing similar positive effects —
for the entire ecosystem have not been accomplished. Since these •
agronomic and natural ecosystems are often physically proximal to one ~
another, further research is needed on the potential influence of S
compounds to each singly and collectively. •
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4-7
Acute foliar in j ury occurs following high-dose exposures and the
rapid absorption of a toxic dose of SC>2 results at first in
marginal and interveinal areas having a dark-green, watersoaked
appearance. After desiccation and bleaching of tissues, the affected
areas become light ivory to white in most broadleaf plants. Some
species show darker colours (brown or red), but there is characteris-
tically an exact line of demarcation between symptomatic and asympto-
matic portions of leaf tissues. Bifacial necrosis is common. In
monocotyledons (e.g., corn, grasses) foliar injury occurs at the tips
and in strips along the veins (Malhotra and Blauel 1980; USEPA
1976).
Plant injury that is visible but does not involve collapse and
necrosis of tissues is termed chronic injury. This type of visible
injury is usually the result of variable fumigations consisting of
both short-term, high-concentration or long-term, low-concentration
exposures to
In broadleaf plants, chronic injury is usually expressed in tissues
found between the veins, with various forms of chlorosis predomi-
nating. Chlorotic spots or chlorotic mottle may persist following
exposure or may subside and disappear following pollutant removal or
as a result of changing environmental conditions (Jacobson and Hill
1970).
The presence of acute or chronic foliar injury is not necessarily
associated with growth or yield effects. Furthermore, when present,
the degree of foliar injury may not always be a reliable indicator of
subsequent growth or yield effects. The uniformity of exposure to
even the low doses of 862 experienced by crops growing under field
conditions presents difficulty in measuring 'treatment1 effects due
to the lack of a set of control (nonpollutant exposed) plants.
Artificial systems must therefore be used under more controlled
laboratory and field situations. The more ubiquitous exposure to
known phytotoxic concentrations of 03 must also be recognized and
singly evaluated.
Yield effects in the absence of foliar symptoms have been reported
for soybeans by Sprugel et al. (1980) and Reinert and Weber (1980)
under field conditions using a zonal air pollution delivery system
and using chamber exposures. Both reports, however, used doses more
typical of point sources of emission and would therefore not be
considered comparable to regional conditions of exposure. No studies
consider all the potential variables that can effect plant response.
This is not a possibility for a single study and is especially true
for field studies (which are most relevant) where many environmental
variables cannot be controlled. From the data available, we can
conclude that growth and yield effects are not necessarily related to
foliar injury. Depending upon the plant affected, the environmental
conditions, and the pollutant exposure conditions, one may observe
yield effects without injury, injury without yield effects or more
direct correlations between injury and yield.
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4-8
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The primary focus of dose-response studies should be to develop
useful generalizations of the relationship between meaningful
parameters of plant response and measurable indices of exposure dose. •
The relationship between exposure dose and the amount of pollutant I
entering the plant may be significantly influenced by environmental
factors controlling the rate of pollutant flux into plant leaf _
tissues (see Figure 4-3). The dose of 862 must be considered in •
relation to known concentrations under field conditions since both *
the regionally expected dose and the phytotoxicity of S02 are
comparatively low (e.g., ozone dose and phytotoxicity are relatively 4
high). •
The role of short-term fluctuations in S02 may be of particular m
importance in areas proximal to point sources of SC>2 (Mclaughlin V
and Lee 1974). Here concentrations may fluctuate widely during
exposure and damage to vegetation may be closely associated with
short-term averages (1 hr) or even peak concentrations. McLaughlin •
et al. (1979) studied the effects of varying the peak to mean S02 ™
concentration ratio on kidney beans in short-term (3 hr) exposures to
SC>2. They found that increasing the peakrmean ratio from 1.0 B
(steady state exposure at 0.5 ppm for 3 hr) to 2.0 (3 hr exposure |
with peak = 1.0 ppm) did not alter post fumigation photosynthetic
depression. However, further increasing the ratio to 6.0 (1 hr _
exposure with peak =2.0 ppm) tripled the post fumigation •
photosynthetic depression. Total dose delivered in the three
exposures was 1.5, 1.8, and 1.1 ppm respectively. Clearly the
quantity of S02 to which the plants are exposed may have a very I
different effective potential as the kinetics of the exposure are ^
changed.
Data on S02 effects on plant growth and yield in most cases provide f
the most relevant basis for studying dose-response relationships. As
a whole-plant measurement, plant productivity is an integrative ^
parameter which considers the net effect of multiple factors over •
time. Productivity data are presently available for a wide range of ™
species under a broad range of experimental conditions. Because
results would not be expected to be closely comparable across these •
sometimes divergent experimental techniques, data have been tabulated flj
separately for only controlled field exposures (Tables 4-la and
4-lb). •
Relatively few crops of economic importance have been studied under
field conditions utilizing various field exposure systems. Of the ^
seven "studies" reviewed in Tables 4-la and 4-lb, dose exposure to •
induce a yield effect was 0.09 ppm S02 for 4.2 hr average ^
fumigation period on 18 days scattered from July 19 through August 27
of the soybean growing season (Sprugel et al. 1980). Five studies •
indicated no effect following various exposure regimes, and one study |
(Neely and Wilhour pers. comm.) reported increased yields (27% and
8%) of winter wheat cv. Yamhill following exposure dose of 0.03 and ^
0.06 ppm S02 for 24 hr/day for the entire growing season, •
respectively.
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4-9
POLLUTANT
CONCENTRATION
NUMBER OF
EXPOSURES
CLIMATIC FACTORS
EDAPHIC FACTORS
BIOTIC FACTORS
PLANT RECEPTOR
MECHANISM OF ACTION
DURATION OF
'EACH EXPOSURE
-GENETIC MAKEUP
STAGE OF PLANT
DEVELOPMENT
EFFECTS
ACUTE
CHRONIC SUBTLE
Figure 4-3. Conceptual model of the factors involved in air
pollution effects (dose-response) on vegetation
(Heck and Brandt 1977).
-------
4-10
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4-12
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Tables 4-la and 4-lb also reviewed a large number of studies which
were conducted using various greenhouse or exposure chamber
techniques and exposure of agronomic or horticultural crop plants. f|
Conclusions indicated difficulty in determining the significance of |
results of such studies in relation to actual similar fumigations
under field conditions. Doses used for exposure treatments were m
usually considered to be in excess of expected doses for ambient I
field exposures. Acute foliar effects have not been reported in
long-term studies using less than 0.15 ppm S02 for 24 hr/day for 7
days. •
In greenhouse experiments conducted in England using ryegrasses,
yield losses were measured following long-term exposure to low levels fl
of SC>2. In one study (Bell and Clough 1973), perennial ryegrass |
experienced a 52% reduction in dry weight after exposure to a mean
concentration of 0.067 ppm S02 over a 26~wk period. At the end of M
the study the plants were smaller and chlorotic in comparison to the H
control plants exposed to air that was purified by both activated
charcoal and an absolute filter. In the other study (Crittenden and
Read 1978), shoot dryweight of Italian ryegrass was reduced by 30 to •
40% after 8-10 wk of exposure to 0.02 to 0.03 ppm S02, and was •
reduced about 10% after 5-wk exposure to air containing 0.004 to 0.02
ppm S02« The Italian ryegrass plants did not display visible •
symptoms of air pollution injury in either the exposure chamber or |
the control filtered air chamber.
In spite of differences due to exposure regimes, techniques, and •
species, certain generalizations can be made with respect to average
and outer-limit responses of the plants under study. These have been
made in the form of correlations of yield response with total •
exposure dose in part-per-million hours (ppmh). The latter data were W
calculated as the product of exposure time and SC>2 concentration
and transformed to log values. For experiments employing controlled •
exposures under field conditions (Tables 4-la and b), data are ^
graphed in Figure 4-4 (McLaughlin 1980). For the 36 data points
shown, exposure dose ranged from 0.24 to 259 ppmh. No effects on —
yield were detected in any of the six studies at doses _>_ 6 ppmh. •
Yield losses occurred in 26 cases at levels ^_ 6 ppmh, while no
effects and positive effects were noted in two cases each at levels
_> 6 ppmh. A linear regression of yield on dose for all studies •
reporting yield losses showed strong positive correlation (r = 0.75) •
of yield with dose and took the form:
Yield loss = -13.6 + 23.8 (log dose) |
r2 = 0.53 (Significance = >_ 0.001)
This correlation excludes four data points, two with no effects and •
two with positive responses. All were studies with wheat reported by *
Neely and Wilhour (pers. comm.). Data from studies reporting no
effect or a positive effect are however all plotted in Figure 4-4. fl
Calculation of the phytotoxic potential for regional scale S02
exposures involves many assumptions regarding toxic and nontoxic m
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4-13
REGRESSION LINE:
% YIELD LOSS —13.6 + 23.8 (LOG DOSE)
r2*0.53 P>F< 0.001
22 DATA POINTS
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0.1
1.0 10.0
EXPOSURE DOSE(ppmh)
100.0
Figure 4-4.
Regression of yield response vs. transformed dose
(ppnh) for controlled exposures using field chambers
(zero and positive effects excluded from regression
analysis) (after McLaughlin 1980).
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4-14
4.2.1.4 S02 Effects to Forest Vegetation
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components of the total dose to which vegetation is exposed.
Obviously not all, but probably most exposures to S(>2 on a regional
scale are below levels producing phytotoxic reactions. An important H
aspect of evaluating the likelihood that plants will be negatively |
influenced by S02 exposures is the determination of what components
within a plant's total exposure history are phytotoxic. Mclaughlin M
(1980) recently examined USEPA (1978b) data on regional S02 concen- •
tration averages. Using the assumption that only the upper 10% of
all S02 exposure days would have S02 concentrations high enough
to cause stress to vegetation, and that only daylight exposure •
(8 hr/day) during the active growing season (6 mo/yr) would be •
effective, he calculated that the average potentially phytotoxic dose
within designated air quality control regions would range from 0.9 •
ppmh (Region IX) to 5.5. ppmh (Region VIII). Maximum doses (highest Jf
reporting stations within regions) ranged from 2.6 ppmh to 27 ppmh,
thus pointing once again to the potential injury to vegetation grown ^
within smaller areas of high S02 point source emissions. •
I
The effects of S02 on broadleaf tree species and similar types of
native vegetation closely resemble those as described for agronomic •
crops. •
In conifers, acute injury on foliage usually appears as a bright —
orange red tip necrosis on the current-year needles, often with a •
sharp line of demarcation between the injured tips and the normally •
green bases. Occasionally, the injury may occur as bands at the tip,
middle, or base of the needles (Linzon 1972). A
Recently incurred injury is light coloured but later bright orange or
red colours are typical for the banded areas and tips. As needle M
tips die, they become brittle and break or whole needles drop from I
the tree. Pine needles are most sensitive to S02 during the period
of rapid needle elongation but injury may also occur on mature
needles (Davis 1972). •
Chronic effects of S02 in conifers are generally first expressed on
older needles (Linzon 1966). Chlorosis of tissues starting at the •
tips progresses down the needle towards the base (i.e., symptoms ||
progress from the oldest to youngest tissues). Advanced symptoms may
follow, involving reddening of affected tissues. Continued chronic M
injury to perennial foliage of coniferous trees results in premature •
needle abscission, reduced radial and volume growth, and early death
of the trees (Linzon 1978).
Forest trees vary considerably in their sensitivity to S02 doses W
and Jones et al. (1973) evaluated the response of numerous species
growing near point sources in southeastern U.S. (Table 4-2). Visible
symptom expression only occurred on the most sensitive species at
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4-15
TABLE 4-2. SULPHUR DIOXIDE CONCENTRATION CAUSING VISIBLE INJURY TO
VARIOUS SENSITIVITY GROUPING OF VEGETATION3 (Jones et
al. 1973)
Maximum Sensitivity grouping
average
concentration Sensitive Intermediate Resistant
(ppm S02) (ppm S02) (ppm SC^)
Peak 1.0-1.5
1 hr 0.5-1.0
3 hr 0.3-0.6
Ragweeds
Legumes
Blackberry
Southern pines
Red and black oaks
White ash
Sumacs
1.5-2.0 2.0
1.0-2.0 2.0
0.6-0.8 8.0
Maples White oaks
Locust Potato
Sweetgum Upland cotton
Cherry Corn
Elms Dogwood
Tuliptree Peach
Many crop
and garden
species
a Based on observations over a 20-year period of visible injury
occuring on over 120 species growing in the vicinities of coal-
fired plants in the southeastern United States.
-------
4-16
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doses of 0.30 ppm/3 hr thus once again pointing to the smaller area
of source influence on direct foliar injury. Dreisinger and McGovern
(1970) indicated a somewhat similar injury threshold (i.e., 0.26 ppm •
S02/4 hr) for visible foliar injury to the most sensitive vegetation I
to S02, but doses were still above ambient concentrations as
expected on a regional basis. •
A few major investigations of the effects of S02 on tree species
growing under natural conditions have been reported (Dreisinger 1965; ^
Dreisinger and McGovern 1970; Linzon 1971, 1978). These reports •
indicated that a pollution (802) gradient existed within the ™
designated study area near Sudbury, Ontario, and effects correlated
well with this gradient. Chronic effects on forest growth were fl
prominent where S02 air concentrations during the growing season |
averaged 0.017 ppm S02, and were only slight in areas receiving
0.008 ppm S02 (Linzon 1978). In Czechoslovakia, Materna et al. m
(1969) reported the occurrence of moderate chronic injury to foliage •
of spruce trees at Celna, under the influence of an average annual
concentration of S02 at 0.019 ppm. ^
Table 4-3 summarizes the results of tree studies that have utilized •
artificial exposure chamber systems under laboratory conditions.
Only two studies (exposures) used doses close to ambient concentra- •
tions (Houston 1974); however, the use of selected clones of known ^j
sensitivity to S02 hinders further field speculation from this
study. The remainder of the studies presented in Table 4-3 have used M
doses above expected occasional exposures under field conditions. •
Concentrations of 0.25 ppm S02 for 2 hr were required to induce
slight injury to several pine species (Berry 1971), but overall
trends for increasing foliar injury do not follow increasing dose for •
conifers per se. Smith and Davis (1978) exposed several conifers W
(pine, spruce, fir and Douglas fir) to doses of 1.0 ppm S02 for 4
hr or 2.0 ppm S02 for 2 hr and only pines developed necrotic tips •
at the 2.0 ppm dose. Likewise, Keller (1980) found only trends in f
reduced photosynthesis in Norway spruce at S02 doses of 0.05 ppm
S02 for 10 wk exposure with significant effects noted at 0.10 and M
0.20 ppm S02 over the same period. •
4.2.1.5 S02 Effects to Natural Ecosystems I
Ecosystems are basically energy processing systems whose components
have evolved together over a long period of time. They are composed •
of living organisms together with their physical environmental •
conditions. Ecosystems respond to environmental changes or perturba-
tions only through the response of the organisms of which they are
composed (Smith 1980). The living (biotic) and nonliving (abiotic) •
units are linked together by functional interdependence. Processes •
necessary for the existence of all life, the flow of energy and
cycling of nutrients are based on relationships that exist among the •
organisms within the system (Billings 1978; Odum 1971; Smith 1980). |
Because of these relationships, unique attributes emerge when
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4-24
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ecosystems are studied that are not observable when individuals,
populations or communities are studied.
Natural ecosystems are seldom, if ever, exposed to a single air «•
pollutant. Therefore, the responses observed under ambient
conditions cannot conclusively be attributed to a single substance •
such as sulphur dioxide alone. Consideration of low SC>2 doses on a |
regional basis presents even further difficulties in discerning
effects induced by this pollutant. .
Questions relating how sulphur deposition from anthropogenic
emissions is incorporated and distributed by aquatic and terrestrial
ecosystems is not fully resolved. The issue is critical since •
ecosystems subject to excess nutrients or toxic materials do not •
commonly distribute them uniformly throughout the system but rather
preferentially sequester them in specific pools or compartments. In m
addition, sulphur dioxide as a gas can cause injury to the vegetative •
components of specific and local ecosystems so that energy flow and
the cycling of other nutrients as well as sulphur may be disrupted if ^
the pollutant is at sufficient concentrations. •
Specific studies of the more detailed effects of SC>2 on natural
systems have been conducted proximal to point sources of high 862 B
emissions and include studies in the vicinity of the Kaybob gas V
plants (Fox Creek, Alberta) (Winner et al. 1978)5 West Whitecourt gas
plant (Whitecourt, Alberta) (Legge et al. 1976) and the Sudbury, •
Ontario smelter district (Dreisinger and McGovern 1970; Linzon 1971). •
Additionally, a series of designed studies using ariticial sources of
S02 have been conducted in the Montana grasslands (Preston 1979).
The results of these studies, particularly the West Whitecourt and •
Montana grasslands studies, document the usefulness of addressing
ecosystem level responses to S02 from a multidisciplinary approach •
incorporating investigations of physiology, autecology, synecology, |
geochemistry, meteorology and modelling. The results confirm that
producers are sensitive to direct S02 effects as evidenced by mm
S02~associated changes in cell biochemistry, physiology, growth, I
development, survival, fecundity, and community composition. Such
responses are not unexpected. An equally important point of
agreement among the different research efforts is the potential for •
ecological modification resulting from either direct S02 effects on •
nonproducer species or direct changes in habitat parameters, which in
turn affect an organism's performance. Changes in biogeochemistry, •
particularly in the soil compartment, are notably responsive to •
low-dose S02 exposures. A major conclusion of the Montana
grasslands studies indicated that at S02 levels above 0.02 ppm (52 _
yg/nH), induced changes occur in the performance of producers, •
consumers, and decomposers. Many of the responses are individually *
small, but collectively over time they gradually modified the
structure and function of the grasslands. The significance of these B
changes to the long-term persistence of the ecosystem remains |
controversial (Preston 1979).
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4-25
Direct effects of SC>2 on individuals within natural plant
communities are most noted within the lichens. Sulphur pollution not
only has caused the depletion of lichen vegetation in certain areas,
but also has resulted in changes in the distribution of different
species (Hawksworth et al. 1973). Epiphytic lichen communities have
been mapped within several regions of North America. In a rural area
of Ohio surrounding a coal-consuming power station (emitting 1025
tons S02/day), the distribution of two corticolose lichens,
Parmelia caperata and I>. ruderta, was markedly affected by elevated
S02 levels (Showman 1975). In regions experiencing an annual S02
average exceeding 0.020 ppm, both species were absent. The
distribution of more resistant lichens was not noticeably affected
until S02 levels exceeded 0.025 ppm (annual average). Somewhat
lower levels were projected by LeBlanc and Rao (1973) to effect the
ability of sensitive lichen species to survive and reproduce; acute
and chronic symptoms of S02 toxicity in epiphytic lichens occurred
when annual averages of SC>2 exceeded 0.03 and 0.006-0.03 ppm
respectively.
The network of biotic-abiotic interactions, which is characteristic
of managed and natural ecosystems, leads to the hypothesis that S02
effects on producers must have repercussions to other trophic levels.
Demonstration of such responses, however, is difficult experimen-
tally, and an accurate assessment of the specific importance of S02
in eliciting these responses is complicated by the often complex
relationships between producers, consumers, and decomposers.
More subtle effects may occur in areas of low S02 (0.05 ppm annual
average) deposition by shifts in soil microfloral populations thus
further influencing plant rhizopheres leading to subsequent ecosystem
alterations (Legge et al. 1976; Wainwright 1979).
Induced changes in natural ecosystems should not be evaluated on a
positive or negative basis. Change as induced by anthropogenic
sources of 862 must be considered as an alteration of natural
processes. For example, natural ecosystems evolved on sulphur-
deficient soils have done so within the imposed constraints per se.
Although atmospherically derived sulphur may not be sufficient to
cause injury, the prolonged input of sulphur may relax the
constraints of a limited sulphur supply thereby inducing shifts in
species composition.
4.2.2 Ozone (03)
Ozone air pollution injury was first reported by Richards et al.
(1958) and during the subsequent years a diverse array of visible
injury symptoms was described on a wide variety of crop, ornamental
and native vegetation. Numerous review chapters and journal articles
contain detailed descriptions of these symptoms (Brandt and Heck
1968; Hill et al. 1970; NAS 1977; USEPA 1978a). Characteristics of
the injury symptoms and extent of injury are influenced by climatic
-------
4-26
4.2.2.1 03 Effects to Agricultural Crops
I
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and edaphic conditions, genetic variability, characteristics of the
pollutant dose, and by interactions between the pollutant and other _
air pollutants or other environmental factors (NAS 1977). Injury •
symptoms described by the various researchers have included: ™
bleaching, bifacial necrosis, general chlorosis, chlorotic mottling,
chlorotic streaking, topical necrosis such as "fleck" and "stipple," I
and pigmented leaf tissue (Hill et al. 1970; NAS 1977; USEPA 1978a). •
In addition to the development of visible injury symptoms, exposure
to atmospheric ozone can: (1) suppress photosynthesis; (2) stimulate •
respiration; (3) inhibit carbohydrate transport; (4) change membrane •
properties; (5) alter metabolite concentrations; (6) alter symbiotic
associations; and (7) alter host-parasite interactions. —
Prior to 1970 most 03 research dealt with observed foliar symptoms •
resulting from acute (short-term), artificially controlled,
dose-response studies. In the 1970s, the research approach shifted •
toward chronic (long-term) studies providing a more realistic |
estimate of natural plant response. The results of several such
studies are summarized in Table 4-4. These studies formed the M
foundation for quantification of dose-response relationships that •
provided a more realistic basis for the assessment of losses under
field conditions. A number of assessment techniques (e.g., open-top
chambers, protective sprays) were utilized in several major studies •
designed to pursue this objective. V
The National Crop Loss Assessment Network (NCLAN) (Heck et al. 1982) •
utilized open-top chambers and controlled 03 concentrations. Its f
purpose was to provide standardized crop dose-response data which
could be utilized in the development of reliable regional scale loss m
assessment calculations. •
I
Foliar responses of crops to artificial 63 exposure have been well
documented and used in the development of species and varietal •
sensitivity listings and the preparation of predictive dose-response £
curves (Larsen and Heck 1976; Linzon et al. 1975). However, these
data may not be reliable for estimating the total effect on crop _
productivity (e.g., yield, quality). Most information now indicates •
that the severity of foliar symptoms is not a reliable index of crop ™
growth or yield effects (Reinert 1980) as there is uneven competition
among several sinks that receive photosynthate. Also, compensatory H
responses to ozone can produce rapid recovery from injury (Jacobson |
in press). Studies with soybeans (Tingey et al. 1973), tomatoes
(Oshima et al. 1975) and alfalfa (Tingey and Reinert 1975) all M
support this concept. The exceptions to this general finding are •
cases where the harvested product is the foliage and where foliar
injury development coincides with the rapid growth of the harvested
product (Linzon et al. 1975). •
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4-27
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4-31
Although the adverse effects of 03 exposure on crop yield or
productivity have not been as extensively documented as has been the
case with foliar injury, there are nevertheless numerous reports on
this topic. Any assessment of yield or quality parameters under
field conditions is complicated by the ubiquity of ozone exposure,
the effect of meteorological variables on ozone distribution within
crop canopies, and the difficulty in establishing ozone-free control
plots. Numerous biotic (pathogen, genetics) and abiotic factors
(i.e., RH, light, and soil moisture) within the environment must also
be taken into account. These difficulties have been partially
overcome by recent progress which has been made in the development of
field assessment techniques for plant growth and productivity
(Reinert 1980). These include open-top field chambers, pollutant
exclusion methods, open-air fumigations, ambient air pollutant
gradients and chemical protectants.
Experimental studies with field grown crops have demonstrated yield
reductions in a large number of ozone-sensitive crops: beans
(Heggestad et al. 1980), potatoes (Heggestad 1973), grapes (Thompson
et al. 1969), corn (Heagle et al. 1972) and others (Heggestad 1980;
Jacobson in press; Reinert 1975). In general the studies have shown
that decreased yield of susceptible species occurs with average ozone
concentrations of between 0.05 and 0.1 ppm for 6-8 hr/day during the
growing season (Heck et al. 1977). In a 5-yr study in Maryland
(1972-79), typical yield reductions were 4, 9, 10, 17 and 20%
respectively for field grown (open-top chambers) snap beans, sweet
corn, potatoes, tomatoes and soybeans (Heggestad 1980).
The first report from the NCLAN project (Heck et al. 1982) appears to
provide good agreement with earlier dose-yield response data (Heagle
and Heck 1980) and with yield losses in the various crops as follows:
soybean 10%, peanut 14-17%, a single turnip 7%, head lettuce 53-56%
and red kidney bean 2%. The yield reductions were equated with
seasonal 7 hr/day mean 03 concentrations of 0.06-0.07 ppm compared
to the 0.025 control value. In the earlier study (Heagle and Heck
1980) employing open-top chambers with 03 dispensing capabilities,
an annual U.S. crop loss estimate assuming a seasonal 7 hr/day mean
03 concentation of 0.06 ppm in all crop production areas was
calculated at $3.02 billion (5.6% of the national production). In a
subsequent manuscript Heck (1981) pointed out that it is a weak
assumption that crops in all parts of the United States are in a
sensitive state during much of the growing season and the values
should be reduced by 50%. This would bring the estimate of 63 crop
losses in the U.S. to between $1 billion and $2 billion or 2-4% of
total production assuming all areas were at concentrations of 0.12
ppm for 1 hr. As most sections of the country are above the current
standard, the national losses are probably higher than the above
values (Heck 1981).
There are limitations in assessing 63 impact on crop species, in
that a majority of presently operating 03 monitors in both the U.S.
and Canada are in urban locations. They therefore may not represent
-------
4-32
I
I
levels to which rural vegetation is exposed. However, some indica-
tion of the occurrence of 03 in rural areas along the U.S./Canada
border is given in Table 4-5. The Ontario rural data (Table 4-6) I
have been summarized to provide some indication of the potential for •
adverse crop effects (growing season daytime basis) and can be
compared directly with the 03 data (Table 4-7) for urban locations •
in the National Air Pollution Surveillance Network (NAPS) in Ontario, |
Quebec and New Brunswick.
It is apparent from these urban and rural data that the southern •
portion of the Province of Ontario is most adversely affected by
ozone in Eastern Canada. This finding is corroborated by numerous
reports of ozone-related crop injuries in this area (Cole and Katz •
1966; Curtis et al. 1975; Hofstra et al. 1978; Ormrod et al. 1980) •
and by the absence of any documented injurious effects to sensitive
agronomic or forest species in Quebec or the Maritime provinces.
I
In Ontario the first indication of transboundary ozone movement
across Lake Erie was documented (Mukammal 1960) following extensive •
work on the relationship between the incidence of weather fleck on •
tobacco and meteorological conditions associated with the buildup of
ozone. Since then a number of large-scale meteorological investi-
gations (Anlauf et al. 1975; Yap and Chung 1977) have documented •
these early findings and have shown that high ozone levels generally •
are associated with regional southerly air flows which have passed
over numerous urban and industrialized areas of the U.S. and which, •
as they move across the lower Great Lakes, undergo rapid dispersion •
as they encounter unstable conditions near the northern shore of Lake
Erie. Contributing to these influx patterns are the more localized _
downwind urban effects which can add to the already high background •
levels. ™
In an effort to estimate the severity and extent of plant injury or B
yield loss resulting from exposure to ambient ozone in southern |
Ontario, a summary has been prepared for all major crop species on
the basis of documented research reports of yield or productivity •
losses in Ontario or the northeastern U.S. and on unpublished •
documents by government agencies or university departments working
under assessment mandates or research contracts. On the basis of
these findings and 1980 economic values it is estimated that the •
average annual loss for ozone-sensitive Ontario crops based on 1980 •
economic values is in excess of $20 million (Pearson 1982). An
example of the types of work which were considered in the assessment •
of crop loss is shown for one of the most sensitive species, white |
bean.
In 1961, bronzing and rusting of white bean foliage was reported •
(Clark and Wensley 1961) throughout southwestern Ontario and the
resultant defoliation and pod abortion was estimated to have resulted
in a loss of approximately 600 pounds of beans per acre (45% yield •
loss) in severely affected fields. Following extensive field work in •
1965 and 1967 the disorder was found to be associated with the
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4-33
TABLE 4-5. THE NUMBER OF TIMES IN 1980 and 1981 THAT OZONE
CONCENTRATIONS EXCEEDED THE USEPA STANDARD OF 0.12 ppm
ALONG THE U.S./CANADA BORDER3
Station
Allen Park
Detroit
Detroit
Essexville
Livonia
Macomb Co .
Marquette Co.
Port Huron
Port Huron
Southfield
Warren
Lake Co .
St. Louis Co.
Berlin
Amherst
Erie Co.
Essex Co.
Monroe Co.
Niagara Co.
Niagara Falls
Rochester
Wayne Co .
Berea
Cleveland
Conneaut
Elyria
Elyria
Painesville
Toledo
Toledo
Westlake
Burlington
Burlington
State
MI
MI
MI
MI
MI
MI
MI
MI
MI
MI
MI
MI
MI
NH
NY
NY
NY
NY
NY
NY
NY
NY
OH
OH
OH
OH
OH
OH
OH
OH
OH
VT
VT
1980
1
2
6
1
1
6
0
5
-
0
0
0
-
-
0
-
5
1
5
2
1
2
0
0
1
0
2
1
0
3
0
0
0
1981
1
0
4
-
1
6
0
7
7
0
0
0
0
0
0
1
7
0
1
0
1
1
1
0
2
-
1
-
5
2
0
0
0
Only data from the U.S. counties touching the international
boundary were used. Data were compiled by Rambo and Patent
(pers. comm.). SAROAD data base covers all of calendar year 1980,
but only includes January to September of 1981.
-------
4-34
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4-38
4.2.2.2 03 Effects to Forest Vegetation
I
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occurrence of elevated levels of atmospheric ozone pollution (Weaver
and Jackson 1968). The symptoms first appear sometime between «
flowering and normal senescence, a critical period in the development •
of yield potential. They appear as a bronze-coloured necrotic
stipple which, as it becomes more severe, results in premature leaf
drop and reduced seed set. •
In an effort to assess and compare the annual severity of ozone
injury on sensitive white beans, Ontario government personnel have •
conducted visual assessment surveys throughout the major production •
areas in southern and southwestern Ontario since 1971. These studies
ruled out any varietal resistance and confirmed that the bronzing _
disorder was widespread throughout all the bean production areas •
(Pearson 1980). •
Studies utilizing chemical protectants against ozone injury have fl
helped to provide information on the economic relevance of the ||
bronzing disorder in Ontario. In one case a 13% yield increase was
associated with the reduction in bronzing severity (Curtis et al. M
1975), while in another study, yield increases of up to 36% (27% •
yield reduction) were realized (Hofstra et al. 1978).
In 1977 and 1978 yield increases with antioxidant protection were not •
as high (Toivonen et al. 1980) due to climatic problems. The overall •
response in these years was 16% and 4% increase in yield respectively
due to antioxidant protection. On the basis of these values and
considering the uniformity of cultivar sensitivity, the average
annual loss for this crop was estimated at 12% (Pearson 1982).
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As in the case of agricultural crops, economic evaluation of the I
effect of pollutants on forest productivity is ultimately contingent •
upon the establishment of dose-response relationships. Consideration
must be given to pollutant loadings and then quantitative measure of •
growth-suppression or yield-depression. •
There are different considerations in evaluating the effects of 03 _
and acidic deposition on forest trees than for agricultural crops. •
Most forest tree species are long-lived, perennial plants that are ™
not subjected to fertilization, soil amendments, cultivation,
extensive pest control or other cultural practices that agricultural •
crops receive. Their size also precludes pollutant exclusion |
(chambers) studies or protective sprays limiting the assessment of
growth or productivity losses to visual observations of growth «
characteristics. This must then be related to ozone dose information •
(i.e., pollution gradients) where available.
In general, many tree species indigenous to North America are •
classified as susceptible to 03 damage (Davis and Wilhour 1976; '
Skelly 1980). Direct injury to tree foliage by 03 has been
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demonstrated repeatedly in experiment situations (Table 4-4), and in
nature as well. Concentrations of 03, at least in some forested
areas, are sufficient to cause injury (Miller and McBride 1975;
Skelly 1980). These effects of 03 can alter the productivity,
successional patterns, and species composition of forests (Smith
1980) and enhance activity of insect pests and some diseases
(Woodwell 1970).
The current status concerning 03~induced effects on Temperate and
Mediterranean forest tree species, communities and ecosystems has
been summarized (Skelly 1980). It is possible that primary
productivity, energy resource flow patterns, biogeochemical patterns
and species successional patterns may all be challenged by oxidant
air pollution.
4.2.3 Acidic Deposition
Various types of injury listed below may result from direct exposure
of plants to acidic deposition (Cowling 1979; Cowling and Dochinger
1980; Tamm and Cowling 1977):
1) Damage to protective surface structure such as
cuticle;
2) Interference with normal functions of guard cells;
3) Poisoning of plant cells, after diffusion of acidic
substances through stomata or cuticle;
4) Disturbance of normal metabolism or growth processes,
without necrosis of plant cells;
5) Alteration of leaf- and root-exudation processes;
6) Interference with reproductive processes;
7) Synergistic interaction with other environmental
stress factors;
8) Accelerated leaching of substances from foliar
organs;
9) Increased susceptibility to drought and other
environmental stress factors;
10) Alteration of symbiotic associations; and
11) Alteration of host-parasite interactions.
In contrast to results with 63, experimental studies with simulated
acidic deposition have produced both positive and negative results
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4-40
4.2.3.1 Acidic Deposition Effects to Agricultural Crops
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(Jacobson 1980; Shriner 1978). Increases and decreases in yield, as
well as no significant effects, have been found. These results
depend upon concentrations of acids, plant species and cultivars, I
pattern and timing of rain applications, and soil, environmental, and I
cultural conditions (Irving and Miller 1980; Lee et al. 1980). Each
species may thus have unique patterns of physiological and genetic •
responses to the potentially beneficial and detrimental components of I
acidic deposition.
I
Experimental studies with plants grown under controlled (or I
semicontrolled) conditions have demonstrated that visible foliar •
symptoms can be produced on certain crops, when pH of applied
simulated rain is 3.5 or less (Table 4-8). Field-grown plants may be •
less susceptible to the development of foliar symptoms than plants I
grown under controlled or semicontrolled conditions (Jacobson 1980;
Shriner 1978). Further, as with 03 and SC>2, foliar symptoms may _
not correlate closely with yield reductions (Lee et al. 1980). •
However, recent evidence suggests that generalizations concerning •
effects on crops from experiments with 03 alone or with acidic
deposition alone, may underestimate the interactive effects of •
sequential exposures to these two pollutants (Jacobson et al. 1980). |
Further research is needed to determine if acidic deposition enhances
the likelihood of actual yield reductions in areas also experiencing H
repeated exposures to elevated concentrations of 03. I
In studies with soils and in studies on aquatic systems focus has
often been on relationships with mean annual deposition rates. •
Characteristics of individual rain events may have greater •
significance in producing direct effects on agricultural crops than
average annual rates. Although annual pH values of rain are as low •
as 4.0 in eastern North America, concentrations of H+ ions (and |
30^2- an(j N03~ ions) may be ten times greater than average
during individual events. The one (or several) most acidic event(s) •
of a growing season may have greater significance for production of •
direct effects on annual crops than average deposition rates.
The potential for crop damage in the field from acidic deposition is I
further amplified substantially by agricultural practices. Economic •
constraints in any given area and year tend to result in the exposure
of extensive areas of a given crop in a relatively uniform state of •
plant development. The onset of the cycle of flowering physiology, •
pollen dispersal and fertilization, and photosynthetic partitioning,
could all be potentially susceptible to extensive damage over vast
areas. •
To evaluate the economic cost of acidic deposition on agricultural
crops, answers to several questions are needed. Which crops are •
actually benefited by components of acidic deposition? Which crops |
are most susceptible to reductions in yield by exposure to acidic
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4-41
TABLE 4-8. REPRESENTATIVE TOLERANCE LIMITS OF SELECTED PLANTS TO SIMULATED ACID
PRECIPITATION3
Plant Species
Birch
Wi 1 low herb
Scots pine
Mosses
Lichens
Sunflower, bean
Hardwoods
Rad i sh
Beet
Carrot
Mustard greens
Sp i nach
Swiss chard
Tobacco
Lettuce
Cau 1 i f 1 ower
Brocco 1 i
Cabbage
Brocco 1 i
Potatoes
Potatoes
Alfalfa
Kidney beans
Oak
Conifer seed 1 ings
Mosses
Chrysanthemums
Juniper
Yel low birch
Pol lutant
Concentration
pH 2.0 - 2.5
pH 3.0
pH 4.0
pH 2.7
pH 2.5
pH 3.5
pH 4.0
pH 4.0
pH 3.5
pH 3.5
pH 3.0
pH 3.0
pH 3.0
pH 3.5
pH 3.5
pH 3.2
pH 2.0
pH 2.0 - 3.0
pH 3.0
pH 4.0+
pH 2.3 - 3.0
Species
Effect
Foliar lesions
Reduced N
fixation rate
Fol iar damage
Fol iar damage
Fol iar damage
Fo 1 iar damage
Reduced yield
Foliar damage
and reduced
marketabi 1 ity
Fol iar damage
Reduced yield
Fol iar damage
reduced yield
Increased yield
Fol iar damage
Increased yield
Inhibition of
parasitic organisms
Fol iar damage
Desiccation, death
Fo 1 i ar damage
and increased
phosphate uptake
Growth decreased
Fo 1 i ar damage
Reference
Abrahamsen et al . 1976
Denni son et al . 1976
Evans et al . 1977
Haines and Waide 1980
Lee et al . 1980
Shriner 1976
Strifler and Kuehn 1976
Teigen et al . 1976
Tukey 1980
Wood and Bormann 1976
The average precipitation pH in eastern North America is currently greater than
or equal to pH 4.0. Individual storm events may have episodes where the pH drops
into the range of pH 3.0 to 4.0.
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4-42
4.2.3.2 Acidic Deposition Effects to Forest Vegetation
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deposition? Unfortunately, only preliminary indications are avail-
able in response to these questions (Lee et al. 1980). Accordingly,
the dose-response function needs to be provided with many more I
quantitative dose descriptions that relate to yield effects under B
actual growing conditions. Information on the influence of other
parameters on these dose-response functions also needs to be •
provided. These factors include patterns of rainfall occurring as |
they interact with stage of crop development, soil nutrient and water
supplies, and deposition of particulate matter from the atmosphere. _
Further clarification also is needed of the possible modifying •
influence of NC>3~ and S0^~ as nutrients in leaf tissue in response ™
to acidic rainfall events. Finally, the critical factors determining
plant susceptibility, expressed as yield reductions, need further I
definition to enhance extrapolation from a few of the most economi- |
cally important crop species and cultivars to describe the response
of the entire ecosystem. •
When this information is provided, it may then be possible to make
reasonable and reliable estimates of the economic impact of acidic
deposition on agricultural productivity. I
I
Effects of acidic deposition on forest trees involves several
considerations differing from those relating to agricultural crops. K
Trees are perennial plants with long lifetimes. Thus, there is •
greater concern with the cumulative impact or repeated exposures to
acidic deposition. Furthermore, forests are usually in areas where
soil nutrient supplies are limited, and are generally not supplied I
with fertilizers or lime. Forests present large surface areas for I
interception of gaseous and particulate pollutants from the atmos-
phere, and these substances eventually move to the soil. Finally, •
the composition of precipitation as it passes through the forest f
system, the properties of soil, and characteristics of streams and
lakes in watersheds are partially affected by the nature, age, and •
condition of forests. Consequently, the effect of acidic deposition •
on forests could also have important secondary impacts which are ™
initiated by direct effects on trees.
The historic pattern of forest growth as revealed in the growth •
rings may show "direct" evidence of the effects of acidic deposition.
Based on substantial analysis of growth rings of Scots pine and •
Norway spruce trees that grow in spatially-intermixed "more I
susceptible" and "less susceptible" regions in south Sweden, Jonsson
and Sundberg (1972) concluded that "acidification cannot be excluded
as a possible cause of the poorer growth development, and may be •
expected to have had an unfavourable effect on growth within the more •
susceptible regions." This is a controversial study because other
Scandinavian researchers have not been able to uncover similar •
trends. For example, in a large study in Norway, Strand (1980) was |
unable to "find definite evidence that acidic deposition has had an
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4-43
effect on the growth of the trees". Studies of a similar type in
North America have been limited in scope. Cogbill (1976), having
examined historic patterns of growth rings in two forest stands (one
a beech-birch-maple woods in New Hampshire and the other a spruce
woods in Tennessee) observed that "no regional, synchronized decrease
in radial increment was evident in the two mature stands studied."
However, Johnson et al. (1981) noted both an abnormal decrease in
growth of pitch pine on the New Jersey pine barrens, and a strong
statistical relationship between stream pH (an index of precipitation
pH) and growth. The relationship of these findings to other possible
incitants (i.e., disease, insects, ozone) should be more fully
explored.
Experimental evidence from studies of the action of simulated acidic
deposition on tree parts does indicate that under regimes of high
acid dosing, direct damage (i.e., foliar lesions) can be produced
(Table 4-8).
A potential impact of acidic deposition may occur indirectly through
the soil and may become involved in the complex natural circulation
of elements upon which forest vegetation depends, (i.e., the nutrient
or biogeochemical cycle). Rodin and Bazilevich (1967) describe this
cycle of elements as "the uptake of elements from the soil and the
atmosphere by living organisms, biosynthesis involving the formation
of new complex compounds, and the return of elements to the soil and
atmosphere with the annual return of part of the organic matter or
with the death of the organisms." Interrelationships in the cycle
are such that a change in one part of the system, if not counter-
acted, could ultimately produce changes throughout.
Generally, forests are relegated to soils which are of low fertility
or, for some other reason, unsuited for agricultural use. In
contrast to agricultural practice, amendments (i.e., fertilizers or
lime) are rarely used in forestry practice.
Deficiencies of nitrogen (N) are common in forests of the temperate
and boreal regions. Appreciable responses to N-fertilizer have been
reported frequently, particularly for conifers on upland sites in
both the acidic deposition zone of eastern Canada (Foster and
Morrison 1981), and in Scandinavia (Malm and Holler 1975; Moller
1972) . In a small number of fertilizer field trials carried out with
conifers in Canadian forests, phosphorus (P), potassium (K), calcium
(Ca) or magnesium (Mg) fertilizers did appear to elicit responses,
though only when demand for N was first met (Foster and Morrison
1981; Morrison et al. 1977a,b).
Growth of red pine and other conifers has been shown to be limited by
K and Mg deficiency in restricted areas of New York State (Heiberg
and White 1951; Leaf 1968, 1970; Stone 1953), and Quebec (Gagnon
1965; Lafond 1958; Swan 1962).
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4-44
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The generally-held association of base-rich with more fertile soils
and base-poor with less fertile soils (well-demonstrated in agricul-
tural situations) has been investigated with forest species and soils •
in only a limited number of instances. Pawluk and Arneman (1961) I
associated better growth of jack pine on sites in Minnesota and
Wisconsin with several soil factors which could be considered acidic
deposition sensitive, including cation exchange capacity (CEC), I
exchangeable K and percent base saturation. Also, in northern B
Ontario Chrosciewicz (1963) associated better growth of jack pine
with soils rich in basic minerals (and presumably richer in
exchangeable bases). Hoyle and Mader (1964) noted a high degree of
correlation between Ca content in foliage and height growth of red
pine in western Massachusetts. Lowry (1972) across a wide range of mm
sites in eastern Canada noted with black spruce relationships between H
site index and foliage content (including N, P, Ca, and to a lesser
extent Mg concentrations).
Studies of forest soils (Lea et al. 1979) indicate that Ca and Mg •
levels can be leached following applications of acidic deposition
simulants. Leaching of these elements from forest soils, as a result
of high S04 mobility (Mellitor and Raynal 1981), may lead to
a chronic decrease in nutrient status of certain soils.
Since nutrient availability is a significant growth-limiting factor •
for many forest ecosystems, the concern is that acidic deposition *
will interfere with uptake and cycling of various elements. First,
acidic deposition may promote increased leaching of essential foliar I
constituents (e.g., K, Ca and Mg) as a function of both acid- •
related surface disintegration and mass exchange by HT1" ions.
Both wet and dry deposition undergo chemical alteration directly on •
the surface of the leaves and indirectly within the cellular tissue.
The nature of the leachate or throughfall depends upon plant _
characteristics such as tree species, leaf morphology, stand I
characteristics (e.g., age and stocking), and site conditions •
(e.g., precipitation rate, distribution and chemical composition).
Input/output analyses and element budgets with particular reference H
to acidic deposition, have been described by various authors (Lakhani |
and Miller 1980; Mayer and Ulrich 1980; Tukey 1980). Generalizations
are difficult, because of the wide range of environmental (i.e., •
soil, water, and climate) conditions. I
Not all elements are leached equally and although all plant parts can
be leached, young leaves are less suceptible to leaching than mature •
foliage (Tukey 1980). Some elements (e.g., K) leach readily from •
both living and dead parts, while others (e.g., Ca) leach more
slowly. •
Some researchers have found that throughfall from deciduous forests
exhibit increased pH and higher Ca and Mg concentrations when •
compared to the incident precipitation. In other instances the I
opposite has been found. In studies of two hardwood species (i.e. ,
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4-45
sugar maple and red alder), little difference in throughfall
chemistry was reported (Lee and Weber 1980). Stemflow from birch
species shows increased acidity, relative to the incident precipita-
tion (Abrahamsen et al. 1977). Beneath coniferous canopies, through-
fall pH generally decreases relative to precipitation in open areas
even though concentrations of Ca and Mg as well as many other
dissolved ions may increase (Horntvedt and Joranger 1976). This ion
enrichment is due to both washout of dry deposits and canopy
leaching. It has been reported that 90% and 70% of the H+ in
precipitation was retained in the forest canopy in New Hampshire
northern hardwood (Hornbeck et al. 1977) and Washington Douglas-fir
(Cole and Johnson 1977) forests, respectively. Leaching of low
molecular weight organic acids from the canopy may decrease the pH of
throughfall (Hoffman et al. 1980).
Spruce canopies may filter dry pollutants from the atmosphere better
than deciduous canopies. This cleansing action is partially
attributed to the presence of spruce needles throughout the winter,
during which S02 is dissolved in water films adhering to their
surfaces. Subsequent removal of these deposits accounts for part of
the difference in chemical composition of the throughfall.
In summary, several processes may be affected when rainfall passes
through a forest canopy. Substances residing on and in foliage are
removed. These processes occur with both acidic and nonacidic
deposition. Certain elements are leached more rapidly than others,
especially when rainfall is acidic. There are also differences
between species and stages of leaf development in rates of leaching.
Leaching results in a marked change in the chemistry of precipitation
before it reaches the soil. Dry deposits removed from leaf surfaces
and substances lost from foliar tissues may neutralize or enhance
acidity and the concentration of inorganic substances may increase
considerably. More rapid transfer of elements to the soil provides
opportunities for enhanced uptake and recycling by trees. Moreover,
soil processes may also be affected by these deposits. Several
pathways exist by which changes to precipitation occurring in the
forest canopy can affect the chemistry of water transported through
the terrestrial ecosystem and into streams and lakes. These are
discussed further in other sections of this report.
Acidic deposition may affect health and/or productivity of forest or
other vegetation through indirect channels, or through effects on
nutrition. Research efforts are just beginning to evaluate the
possible role of acidic deposition in the predisposition of trees to
disease infection and insect attacks. Further, the behaviour of
plant litter and soil-occurring facultative saprobes, which may
exhibit plant pathogenic tendencies under acidification, requires
evaluation.
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4-46
4.2.4 Pollutant Combinations
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For much of the northeast and midwest sections of the United States •
where acidic rainfall events and low dose SC>2 trends have been |
recorded, ozone air pollution also occurs on a concomitant basis.
Sulphur dioxide, NOX, and particulate emissions may be of "local" •
importance to vegetation, but mesoscale background concentrations of
these pollutants are well below known thresholds for inducement of
direct vegetation effects. From these background concentrations, •
long-term accumulation by plants of sulphates and nitrates and the •
related potentially beneficial or detrimental effects are poorly
defined. •
Extrapolation from results of single pollutant effects on vegetation
under ambient field conditions must be approached with caution. ^
Reactions in pollutant combinations may be additive (sum of effects), I
less than additive (antagonistic), or more than additive ™
(synergistic). In addition to pollutant combinations under
controlled conditions, the interaction of constantly changing •
environmental factors and fluctuating pollutant doses must be further •
evaluated before a conclusive statement of the importance of such
interactions can be made. Reinert (1975) and Reinert et al. (1975)
have prepared the most recent reviews of this area of investigation.
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4.2.4.1 S02 - 03 Effects I
The most frequently occurring pollutant combination of significance
to plant life must be considered as 03 and S02• However, few I
studies have utilized doses which would be considered as even close |
to ambient except as they pertain to areas affected by point sources
of emission of S02- Studies using combinations of 03 and S(>2 •
are presented in Table 4-9. As indicated, only the study of Houston •
(1974) used doses of SC>2 approaching regional expected averages.
He used mixtures of S02 and 03 in doses to simulate actual field
conditions and reported that even the lowest concentrations of 03 fl
(0.05 ppm) and S02 (0.05 ppm) for 6 hr in mixture caused more H
serious damage than that resulting from either pollutant alone at
similar concentrations. Studies by Tingey et al. (1971a,b, 1973), •
Tingey and Reinert (1975), and Neely et al. (1977) used doses |
reasonably expected in smaller areas such as the Ohio Valley (Mueller
et al. 1980). Doses used in other studies used less realistic doses _
for either S02 or 03 and the results are of little value in •
estimating field effects on a regional basis. —
A recent study by Reich et al. (1982) utilized a linear gradient •
field exposure system of S02 and 63 over soybeans exposed during •
pod fill. Low dose exposure combinations averaged S02 at 0.040 ppm
and 03 at 0.034 ppm for 5.5 hr per day for 12 days. Yield •
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4-47
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4-48
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4-50
4.2.4.3 S02 - 03 - Acidic Deposition Effects
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expressed as dry mass/seed was 9% less than that of plants exposed to
ambient air; higher doses reduced yield by 10% and 15%. _
4.2.4.2 S02 - N02 Effects
The possibility of adverse effects occurring on plant life due to the |
interaction between atmospheric S02 and N02 needs consideration.
Tingey et al. (1971a) demonstrated experimentally that a gaseous m
mixture of 0.10 ppm S02 and 0.10 ppm N02 caused synergistic •
effects with more than 5% leaf injury being induced on 5 of 6 plant
species treated in 4-hr exposure periods. The symptoms of injury
produced by a mixture of S02 and N02 can resemble those caused by •
ozone which may make diagnoses in the field difficult. *
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The currently available literature concerning the interactive effects _
of acidic deposition and gaseous air pollutants on terrestrial •
vegetation is extremely limited, consisting of only three separate ™
studies. Shriner (1978) examined the interaction of acidic
deposition and S02 or 03 on red kidney bean (Phaseolus vulgaris) M
under greenhouse conditions. Treatments with simulated rain at pH •
4.0 and multiple 63 exposures resulted in a significant reduction
in foliage dry weight. Simulated precipitation and sulphur dioxide m
in combination did not affect either photosynthesis or biomass •
production. Troiano et al. (1981) exposed two cultivars of soybean
to ambient photochemical oxidant and simulated rain at pH 4.0, 3.4, _
and 2.8 in a field chamber system. The interactive effects of •
oxidant and acidic deposition were inconclusive with seed germination ™
being greater in plants grown in the absence of oxidant at each
acidity level. Irving and Miller (1980) also examined the response 4
of field-grown soybeans to simulated acidic deposition at pH 5.3 and P
3.1 in combination with sulphur dioxide and ambient ozone concentra-
tions. No interactive effects of acid treatments with S02 on m
soybean yield occurred. However, sulphur dioxide alone resulted in a •
substantial yield reduction.
Changes in such things as soil chemical properties nutrient •
recycling resulting from acidic deposition do not occur rapidly. •
After more than a decade of research in Scandinavia, the observed
changes in chemical properties of forest soils that can be attributed M
to acidic deposition still remain undetermined (Overrein et al. |
1980). It is therefore unlikely that interactive effects of acidic
deposition and gaseous pollutants on plants involving changes in soil ^
properties will become evident within a single growing season. I
A physical and chemical potential exists for interaction of various
forms of wet fall and dry fall (including gases and trace metals) at, •
on, or within leaf surfaces. However, very few studies have •
addressed these interactions and the significance of the observed
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4-51
phenomena remain inconclusive (Fuzzi 1978; Gravenhorst et al. 1978;
Penkett et al. 1979).
4.3 EFFECTS OF ACIDIC DEPOSITION ON TERRESTRIAL WILDLIFE
Although direct effects of acidic deposition are not likely,
terrestrial wildlife may be indirectly affected in three ways:
(1) contamination by heavy metals mobilized under acidic conditions;
(2) loss of essential nutritional components from food; and (3)
reduction in food resources.
While the sensitivity of organisms such as plankton and fish to
metals released in acid waters has been established (Baker and
Schofield 1980; Marshall et al. 1981; Muniz and Leivestad 1980),
the potential for accumulation and subsequent effects on terrestrial
animals is less well understood. Metal contamination and reduced
size of roe deer (Capreolus capreolus) antlers from an industrialized
area of Poland has been reported recently (Jop 1979; Sawicka-Kapusta
1978). Acidification and sulphurization of roe deer browse
(Sawicka-Kapusta 1978) was suggested as the cause of the high metal
levels (Jop 1979). Such a means of contamination has been
demonstrated in southeastern Denmark where cadmium and copper in
epiphytic lichens and mosses were compared with those from
northwestern Denmark (Gydesen et al. 1981). Epiphytes from the
southeastern areas of Denmark which received elevated metal
deposition in bulk precipitation showed metal levels 1.5 times
higher on average. The same trend was found in the kidneys of cattle
feeding in these areas. While direct deposition to plant surfaces
may be partially responsible, plant uptake of some metals such as
cadmium increases as soil pH decreases (Andersson and Nilsson 1974),
and high plant metal content is another route of contamination. In
Sweden, moose (Alces alces) closer to sources of anthropogenic
sulphur supported higher tissue levels of cadmium and the body burden
increased with age (Frank et al. 1981; Mattson et al. 1981). The
mechanism of contamination was not explored and could be via
terrestrial or aquatic vegetation.
The availability of essential elements in wildlife nutrition may be
affected by sulphur deposition and soil pH. Selenium, for example,
is an essential element for vertebrates (Stadtman 1977). Selenium
deficient conditions lead to degeneration of major body organs such
as the liver, kidney and heart (Harr 1978; Schwarz and Foltz 1957).
Most importantly from the viewpoint of ranchers, muscular dystrophy
(known as white muscle disease) has been caused by selenium
deficiency and reported in sheep, cattle, swine and horses (Harr
1978; Hidiroglou et al. 1965; Muth et al. 1958). The occurrence of
white muscle disease in North American livestock is correlated to the
concentration of selenium in forage (Allaway and Hodgson 1964).
Lameness and poor growth and reproduction in domestic animals have
resulted from selenium-deficient diets (Harr 1978). In poultry,
edema (abnormal excess accumulation of fluid in connective tissue or
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body cavities) has been related to selenium deficiency (Harr 1978;
Patterson et al. 1957).
Many of the soils in the temperate region of eastern North America •
are low in selenium and hence produce forages which contain low
selenium concentrations, frequently less than the 0.1 ppm minimum •
level for animal health (Kubota et al. 1967; Levesque 1974; Winter •
and Gupta 1979). Although selenium deficiencies in livestock have
been associated with forages grown on soils naturally low in _
selenium, many incidences of deficiency have been attributed to the V
high agricultural use of sulphate fertilizers (Allaway 1970; Davies 9
and Watkinson 1966). Calves reared on hay grown in Kapuskasing,
Ontario, developed muscular dystrophy due to selenium deficient I
conditions in the soil there (Lessard et al. 1968). flj
Due to the correlation between soil selenium levels and concentra- M
tions in plants grown on these soils (Muth and Allaway 1963), there w
is evidence that wildlife forage plants are similarly selenium
deficient. This was the finding in a study of moose browse plants in
Alaska (Kubota et al. 1970). Moreover, selenium deficiency symptoms M
have been reported for several wildlife species, (e.g., the prong- IP
horn; Antelocapra americana) (Stoszek et al. 1978). Mountain goats
(Oreamnos americanus) from an area where selenium levels in forage
are low and where white muscle disease occurs in livestock, revealed
symptoms of white muscle disease upon being stressed by handling
(Herbert and Cowan 1971). It is suggested that the symptoms in wild —
populations may well be masked by predation (Herbert and Cowan 1971). •
The net effect of selenium deficiency diseases in wildlife would be ™
an increased susceptibility to predation as well as reduced
productivity and survival of young. •
Recent increases in anthropogenic sulphur emissions have caused
concern regarding the influence on selenium availability in •
vegetation. Selenium concentrations in plants in heavily industrial- •
ized Denmark have decreased over the past decades (Gissel-Nielsen
1975). Experimental applications of S02 and SO/ to plants ,—
and soils have demonstrated that selenium levels are depressed by •
both the presence of sulphur and reduced soil pH (Shaw 1981a,b). ™
Because excessive sulphur and sulphate cause uptake of selenium to be
reduced in plants (Davies and Watkinson 1966; Gissel-Nielsen 1973; fi
Shaw 1981a), the impact in areas of low selenium soils could be |
substantial. Furthermore, the solubility of selenium declines with
pH, rendering selenium less available to be taken up by plants in M
acid soils (Geering et al. 1968; Johnson 1975). |
Sulphur and its compounds have a further depressing effect upon
selenium in the animal itself. Excessive sulphur in the diet can •
lead to increased elimination of selenium from the body (Harr 1978), •
compounding deficiency conditions.
9
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4-53
Other essential elements in animal nutrition such as calcium,
magnesium and sodium are similarly released from soils upon acidifi-
cation (Abrahamsen 1980; Rorison 1980; Stuanes 1980). Accordingly,
such elements will be less available for uptake by plants, resulting
in lowered concentrations in plant tissues (Beeson 1941). Soil
acidification similarly causes leaching of phosphorous which, if
reflected in the vegetation (Rorison 1980), could have significant
effects on wildlife nutrition. Lucas and Davis (1961) summarize the
influence of pH on the availability of 12 plant nutrients.
Aside from nutrient content, the availability of food resources may
decline due to acidic deposition affecting the entire food web
including wildlife. For example, caribou (Rangifer sp.) may be
affected due to the sensitivity of lichens to sulphur and acid
compounds (Lechowicz 1981; Puckett 1980; Sundstrom and Hallgren
1973). The importance of lichens in the winter diet of Canadian
caribou herds is well documented (Kelsall 1960, 1968; Thompson and
McCourt 1981). Thompson and McCourt (1981) reported that 67% of the
diet of the Porcupine Caribou Herd of the Yukon consists of lichens.
The George River caribou herd of Nouveau Quebec and Labrador is the
largest in North America (Juniper 1979; Juniper and Mercer 1979;
Mallory 1980) and may rely heavily on the carrying capacity of their
winter range. Much of this area lies in the zone of acidic
deposition (Figure 8-lb). Exposure of the primary caribou lichen
(Cladina stellaris) to simulated acidic deposition with pH 4.0,
reduced maximum photosynthesis by 27% and slowed recovery from
dormancy after wetting by 14% (Lechowicz 1981). These results
suggest that acidic deposition reduces the growth and productivity of
this lichen (Lechowicz 1981). The significance of reduced lichen
productivity to the population dynamics of these caribou herds is
uncertain, because the degree to which they are food-limited is
unknown.
Another example of potential food loss involves herbaceous ground
cover. Trees have tap roots in deep soil layers that are less
susceptible to acidification, while plants draw their moisture and
nutrients from the upper layers of soil making them more exposed to
the effects of acidic deposition (Clark and Fischer 1981).
Application of sulphuric acid in quantities corresponding to
100 kg/ha.yr killed much of the ground vegetation consisting mainly
of mosses, lichens, and a species of dwarf shrub (Tamm et al. 1977).
Therefore animals which feed on such vegetation may be affected by
food loss.
4.4 EFFECTS ON SOIL
Soils vary widely with respect to their properties (i.e., physical,
biological, chemical and mineralogical), support different
vegetation communities, are subjected to different cultural
practices, are situated in different climatic zones, and are exposed
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to a broad spectrum of acid loadings making it difficult to
generalize from findings indicated in this report. Further, there
are various offsetting mechanisms, influencing effects of increased •
precipitation acidity which vary with soil properties, vegetation •
types, climatic regimes and cultural practices. Water moves through
soils by uniform capillarity and gravitational processes. Also, A
considerable moisture flow may be directed overland or may be <|
channelized in the soil in root channels reducing the opportunity for
equilibration. A high stone content can concentrate leaching effects ~
to a smaller soil volume than in nonstony soils. Thus, theoretical •
calculations have to take into account particular ir± situ character- ™
istics.
In the discussion which follows, the documented and hypothesized 9
effects of acidic deposition on soils are described under the
following headings: •
1. Effects on Soil pH and Acidity.
2. Impact on Mobile Anion Availability, Base Leaching, and Cation V
Availability. ™
3. Influence on Soil Biota and Decomposition/Mineralization fl
Activities. I
4. Influence on Phosphorus Availability. <^fl
5. Effects on Trace Element and Heavy Metal Mobilization and
Toxicity. —
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4.4.1 Effects on Soil pH and Acidity
acidifying sulphate fertilizers brings about appreciable soil
acidification, along with other changes in soil chemical and •
biological properties (Glass et al. 1980). The more striking of •
these changes are reductions in exchangeable bases, increases in
soluble aluminum and manganese levels, shifts in optimum conditions ^
for bacteria and mycorrhizal fungi, and reductions in soil micro- •
faunal populations. Some of these undesirable changes have also been •
shown to occur in the proximity of strong point emitters of sulphur
dioxide (Freedman and Hutchinson 1980; Nyborg et al. 1976; Strojan •
1978), so concern is well-founded that the range of soil changes |
outlined in Table 4-10 could occur to a greater or lesser extent over
more widespread geographical areas. M
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The process of soil acidification primarily involves the replacement
of exchangeable basic cations (Ca, Mg, K, Na, NH^"1") by H+ and,
at lower pH ranges, Al^"*" ions. The chemistry of soil acidification •
is relatively well understood, at least in states other than strong •
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TABLE 4-10. ACIDITY RELATED REACTIONS INFLUENCING AVAILABILITY
OF SEVERAL ELEMENTS
Element(s) Type of Reaction
N Chiefly biological -
biochemical; nitrifying bacteria
decline with declining pH, thus
ammoniacal-N predominates over
nitrate-N; reduces mineralization.
P Phosphate fixation reactions.
K, Ca, Mg Chiefly mass displacement of
absorbed bases by H and Al^+ ions.
Fe, Mn Chiefly dissolution of hydroxides
in acid solution; organic status,
redox important particularly for Fe.
Al pH regulated solubility of Al-oxy
and hydroxy compounds.
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acid as described by Jenny (1961), Wiklander (1973/74, 1975, 1980a),
Bolt and Bruggenwert (1978), Bache (1980), and Nilsson (1980). In
the most strongly acid soils, there is evidence that aluminum becomes •
very mobile without there being any associated notable change in pH ™
(Cronan and Schofield 1979; Norton et al. 1981; Ouellet 1981; Ulrich
et al. 1980). The common range of pH for soils in humid regions is fl
about pH 5.0 to 7.0, with the preferred range for cultivated soils ||
being pH 6.0 to 7.0. Many forested soils, particularly under
coniferous cover, fall within the range pH 4.5-5.5 in the mineral •
horizons, with surface organic layers commonly exhibiting pHs in the •
range 3.5 to 4.5.
The numbers of field situations where investigators have been able to M
compare present with former soil pH values are extremely limited. 9
However, Linzon and Temple (1980) report a lowering of soil pH in the
brunizolics, but not podzols, of south-central Ontario after 18 years ft
of pollutant deposition. Ulrich (1980b) and his colleagues (Ulrich \|
et al. 1980) working in the more heavily polluted parts of central
Germany report a long-term fluctuation of pH in the surface humus —
under beech and spruce. The pH values do not show a steady decline, Tm
but rather show cyclic variation between 4.2 and 3.8. This parallels ™
deacidification and acidification phases alternating between cooler,
moister summers and warmer, drier ones. From 1969 to 1980 under M
beech, and from 1973 to 1980 under spruce, there were substantial •
increases in the amounts of soil aluminum mobilized. These increases
were associated with the continued entrapment and deposition of acid •
sulphate pollutant. •
Various field and laboratory experiments of a simulation nature have _
also been set up to examine the effects of acidic deposition on soil fl
acidity. Results indicate that artificial acidic deposition at pH<4 ™
can lead to measurable decreases in soil pH (Abrahamsen et al. 1976;
Bjor and Teigen 1980; Stuanes 1980). For example, simulated acidic 4
deposition inputs of pH 4.0 and below to spruce podzol soils in ^
Norway caused soil acidification of the 0, A, and B horizons
(Abrahamsen et al. 1976). In some cases, the soil pH depression over
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4-57
pH, 4.9), carbonic acid contributed approximately twice as much H"1"
to the soil as did precipitation. However, a drop in pH to 4.0
(about 30 times more acid than normal) occurs in the most heavily
impacted areas of eastern North America (Cogbill and Likens 1974) and
results in H+ inputs far in excess of that produced by carbonic
acid. In the more acid soils having a pH of less than 5.5 (e.g.,
podzols developed under coniferous forests), organic acids contribute
significantly to natural soil acidification. It is not as yet known
what role anthropogenic acidification will have in these ecosystems.
Presumably, extremely acid soils will experience the least change in
pH, but changes in the ionic make-up of the soil colloidal complex
and ionic mobility may take place.
Sollins et al. (1980) proposed a comprehensive scheme for calculating
H+ ion budgets in forest ecosystems, based upon measured mass
balances of cations and anions within the nutrient cycles.
Andersson et al. (1980b) used this model to obtain H+ ion budgets
for forest ecosystems in Sweden, West Germany, and Oregon. In the
heavily impacted Soiling site in West Germany, their analysis shows
that atmospheric H+ ion inputs are small (approximately 10%)
compared to net internal flows. Ulrich (1980b), using essentially
the same approach, stressed input-output balances to assess the
long-term net acidification of soils caused by internal compensations
of H+ production and consumption and uptake and mineralization
processes. He also pointed out important spatial considerations
within the soil profile. For example, ammonium mineralization that
consumes hydrogen ions might occur in litter layers while ammonium
that produces hydrogen ions may occur in mineral soil layers at the
same time. Some indication of orders of magnitude of H"1" ion
contribution by softwood versus hardwood forest and their
relationship to anthropogenic loading, were provided (Ulrich 1980b).
Total H+ ion input was determined as about ca 0.81 keq/ha, of which
0.79 keq/ha was considered man-made. A beech canopy generated an
additional ca 0.58 keq/ha and a spruce canopy, an additional 2.28
keq/ha. This evidence suggests that as mean pH of rainfall declines
below pH 4, its contribution to the H+ ion balance is not
insignificant even in comparison to spruce forest H+ ion
production. Thus, the process of podzolization is hastened.
As noted earlier, the adverse effect of soil acidification results
chiefly from the influence of changed pH on other processes (e.g.,
soil biochemical reactions and N availability, organic matter
turnover, mobilization of trace elements, and transformation of clay
minerals) .
4.4.2 Impact on Mobile Anion Availability and Base Leaching
Acidification and soil impoverishment involves the displacement of
basic cations (i.e., K, Ca, Mg, Na) from exchange surfaces, their
replacement by H"1" and Al^+ ions, and the establishment of new
exchange/solution equilibria (Wiklander 1973/74). Under natural
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conditions, two sets of processes seem to be involved: (1) there are
exchange processes whereby H"1" ions displace base cations from the _
exchange surface, and (2) there are the processes whereby the •
exchanged ions are transported within the soil column under the ™
influence of mobile anions (Johnson and Cole 1976, 1977).
_
capacity or CEC and the relative degree of saturation of the CEC with
bases or base saturation). In humid regions, the total permanent and m
pH dependant CEC of a productive soil under cultivation might range I
from 15 to 30 meq/100 g. In the surface horizons this might be
higher and in the subsoil it may be lower. To illustrate this, in
coniferous podzols the CEC of the humus layer may be high while I
beneath it values decrease abruptly with depth. It is presumed that Qi
the loss, particularly of those base cations of nutritive value
(chiefly K, Ca, and Mg), could be accelerated under acidic
deposition, with attendant adverse impacts on forest growth.
I
Various "simulated acidic deposition" leaching experiments are »
described in the literature (Abrahamsen et al. 1976; Abrahamsen and M
Stuanes 1980; Lee and Weber 1980; Morrison 1981; Overrein 1972; *
Roberts et al. 1980; Singh et al . 1980). In some controlled
irrigation experiments, Ca and Mg appear to be the most affected and •
K the least affected (Abrahamsen 1980; Hovland et al. 1980; Ogner •
and Teigen 1981; Wood and Bormann 1976). To some extent, this may
reflect the relative amounts of these cations on soil exchange sites, •
but the rate of increase in K depletion seems to be consistently •
below that for Ca or Mg under acid irrigation as well (Abrahamsen
1980; Ogner and Teigen 1981; Wood and Bormann 1976). The relative ^
lack of response in K may also be due to the greater plant fl
requirements for K, as opposed to Ca or Mg, and possibly also to ~
fixation of K in 2:1 clays.
9
In some cases, the accelerated cation leaching has led to net
depletion of available cations in the rooting zone. Significant
reductions of base saturation percentage were noted in the 0 and A •
horizons in Norwegian spruce podzol soils, following applications of •
simulated acidic deposition with a pH of 3.0 or lower (Abrahamsen
1980). ^
Soil acidification and decreases in base saturation do not always ™
occur concurrently. Under natural soil acidification by humic acids,
production of humus increases CEC, but does not increase the cation B
content (Konova 1966). Soil pH and base saturation will thus |
decrease without a corresponding reduction in exchangeable base
content (Ulrich 1980a). Similarly, with anthropogenic acidification, •
soil pH and base saturation may decrease, with no corresponding net I
nutrient loss. This occurs if the soil is actively adsorbing both
H+ and SO^-, which would increase CEC over time (Johnson and
Cole 1977). In addition, decreases in base saturation and pH in •
soils subjected to leaching losses of base cations can be offset to ™
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some extent by acid-induced increases in soil weathering (Johnson
1980).
Much of the potential impact of atmospheric deposition stems from the
input of the mobile S0^~ anion to soils. Whereas the mobility
of bicarbonate or organic anions may be severely limited in many acid
or clay-rich northern soils, SO^" anions may be very mobile in
these same soils. It has been shown that atmospheric H2S04
inputs overwhelm natural leaching processes, in some New Hampshire
Spodosols, causing perhaps a threefold increase in the natural rate
of cation denudation, and marked increases in the leaching of soluble
inorganic Al. In New Hampshire subalpine coniferous soils, anthro-
pogenic 504^" anions supplied 76% of the electrical charge
balance of the leaching solution, while A1.3+ and H+ were the
dominant cations in solution (Cronan 1980; Cronan et al. 1978; Cronan
and Schofield 1979). In contrast, some soils (chiefly those rich in
Fe- and Al-sesquioxides) exhibit a substantial capacity to adsorb
S0^~, and thus demonstrate a considerable initial resistance
to base leaching by anthropogenic ^864 (Johnson and Cole 1977;
Johnson and Henderson 1979; Morrison 1981; Roberts et al. 1980; Singh
et al. 1980). This generally implies that the effect of acidic
deposition on soil cation leaching is highly dependent upon the
mobility of the anion associated with the acid, whether it be
864^", N03~, or an organic anion (Cronan 1980; Johnson and
Cole 1980; Seip 1980). This is due to the requirement for charge
balance in the soil solution, a necessary condition that precludes
the leaching of cations without associated mobile anions. Soils low
in free Fe and Al, or high in organic matter (the latter appears to
block sulphate adsorption sites, [Johnson et al. 1979, 1980]) are
therefore generally susceptible to leaching by H2S04 (e.g.,
Cronan et al. 1978). Where SO,2- adsorption does occur, (e.g.,
in the highly weathered soils of Tennessee), S accumulation could
initially be beneficial in three ways: (1) prevent cation leaching
by H2S04 by immobilization of the 804^" anion; (2) create
new cation exchange sites; and (3) release OH~ from adsorption
surfaces (Johnson et al. 1981). It follows, however, that once
804^" exchange sites become fully occupied, cation leaching
could commence. On Walker Branch Watershed, 48% of total S to input
accumulates in the soil, whereas only 13% accumulates in vegetation
(Johnson and Henderson 1979; Shriner and Henderson 1978). Along the
same lines, one might expect the N03 in acidic deposition to
contribute to net cation leaching only in those systems where
N03~ is mobile. Because of the N-limited status of many forests,
most N03 tends to be assimilated by plants during the growing
season, thereby not contributing to cation leaching.
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4.4.3 Influence of Soil Biota and Decomposition/Mineralization
Activities
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It has been postulated that atmospheric deposition of strong acids
may adversely affect soil biota and decomposition activities either
directly through soil acidification or indirectly through trace tt
metal mobilization and toxicity. Laboratory experiments and •
observations on soils in close proximity to pollutant sources provide
information on changes which occur in soil biota, as a result of —
increased acidic deposition. Observations indicate decreases in j
total numbers of soil bacteria and actinomycetes, and some relative ™
increase in presence of fungi; although, under conditions of very
high loading, fungi have been reported less abundant. Generally, H
total numbers of enchytraeids have not been affected (except under £
extreme conditions), though differential species responses have been
reported (Abrahamsen et al. 1976, 1977; Alexander 1980; Baath et al. *
1980). '|
The available evidence on the effect of acidity on organic matter
breakdown and soil respiration is not conclusive (Rippon 1980; Tamm ^B>
et al. 1977). However, decomposition experiments suggest that acidic ™
deposition may retard organic matter decomposition. Studies (Baath
et al. 1979, 1980; Francis et al. 1980; Lohm 1980; Tamm et al. 1976) ft
have noted decreased decomposition or carbon mineralization in soils ||
and litter exposed to artificial acidic deposition inputs at pHs
below 3.5 to 3.0. Meanwhile, other studies have shown little or no ^
effect (Abrahamsen 1980; Hovland et al. 1980). Clearly, the results •
are partly dependent on soil type and severity of the simulated ™
acidic deposition treatment.
In some soils, there are indications that acidic deposition may alter •
humic/fulvic acid dynamics. While moderate acidity may aggregate
humic acid particles, it may lead to dissolution and mobilization of •
fulvic acids. In soils like Podzols, which contain appreciable •
quantities of fulvic acid, substantial losses could occur in moderate
acidic leaching.
Besides carbon cycling, there is concern that acidic deposition may *
have adverse effects on N cycling patterns and processes. In this
case, there are actually two potential sides to the issue: (1) the •
possibility that acidic deposition may decrease N mineralization and |
availability, and (2) the possibility that atmospheric inputs of
anthropogenic N compounds may provide a fertilizer effect by increas- M
ing the amount of available nitrogen. Tamm (1976) predicted short- •
term increases in N availability and tree growth, due to net N losses
from ecosystems. In Germany, Ulrich et al. (1980) resampled soils ^
over a 13-year period and showed significant accumulations of N-poor •
organic matter in the forest floor of a 120-year old beech forest. w
This was interpreted as a condition which could lead to internal H+
ion production, immobilization of N, and mobilization of soluble •
Al3+. other studies, by Francis et al. (1980) and Alexander £
(1980) show ammonification and nitrification may decrease markedly in
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soils exposed to artificial rain at pHs approaching 3.0. However,
several studies have demonstrated increased N availability, at least
during the initial stages of H2S04 input (Abrahamsen 1980; Ogner
and Teigen 1981; Roberts et al. 1980), and this has produced minor
growth increases in situations where N is limiting (Abrahamsen 1980;
Tamm and Wiklander 1980; Tveite and Abrahamsen 1980). Whether this
increase in N availability is due to change in microbial activity, or
to the acid catalyzed hydrolysis of labile soil N, is unknown as yet.
Norwegian studies show that both N availability and N03~ leaching
were stimulated by H2S04 inputs-. This strongly suggests that
contrary to earlier predictions (Tamm 1976), nitrification can be
stimulated by acid inputs as well. This has definite negative
long-term implications for forest N and cation status, if NC>3~
production exceeds plant uptake, resulting in net ecosystem N and
cation loss.
4.4.4 Influences on Availability of Phosphorus
Like N, phosphorus (P) is an essential element for plant life. In
soil, P occurs in both inorganic and organic compounds. It is
utilized from the soil solution by plants chiefly, though not
entirely, as the (inorganic) orthophosphate anion. For perennial
plants, including trees, P is assimilated through the intermediary
mechanism of a mycorrhizal root association (Bowen 1973; Fogel 1980;
Hayman 1980). The availability of P to plants is determined to a
large extent by the ionic form in which it is present. In soil
solutions of low pH, available P is present largely as H2P04";
as pH increases, HP042~ predominates. In strongly acid soils,
H2P04~ ions may react with soluble Mn, Al and Fe compounds and
be mostly precipitated as the insoluble and nonavailable metal
hydroxyphosphate (Hsu and Jackson 1960). Also, under conditions of
increasing acidity, H2P04~ tends to react with the insoluble
oxides of Fe, Al and Mn, and in more weathered soils it may become
fixed on silicate clays, through the process of anion exchange.
4.4.5 Effects on Trace Element and Heavy Metal Mobilization and
Toxicity
A further effect of acidic deposition or increased soil acidification
is an increased solubilization of heavy metals in the soil system.
This can arise from the increased solubilization of metals that are
already present in mostly insoluble or nontoxic forms or it may arise
from metals being deposited along with an acidifying pollutant.
Thus, at low concentrations naturally present Mn and Fe serve as
essential nutrient elements for the growth of higher plants and
except in alkaline or calcareous soils are usually present in
adequate available amounts. However, at high concentrations these
metals and Al can cause nutritional imbalance and growth impairment.
Different plant species vary in their susceptibility to heavy metals,
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an example being Al. Barley, sugar beet, corn and alfalfa are very
sensitive, whereas ericaceous shrubs and conifers appear much more
tolerant. Soil characteristics also affect tolerance/susceptibility fl
including soil pH, Ntfy"1" compared with NC>3~ nutrition, Al-exclusion V
processes, Ca and P status, and organic-Al complexation (Foy et al.
1978). Striking examples of the effect of soil pH on the solubility *
of Mn and Al are given in Glass et al. (1980), concentrations rising •
very rapidly in each suite of soils when pH values moved from 5 to
1
In acid forest soils that support highly productive forest in eastern •
Canada it is not unusual to have a pH gradient down the profile of
from 3.0 to 4.6. Associated with such values are exchangeable Ca A
values falling from 2.0 to 0.15 meq/lOOg and exchangeable Al values "p
falling from 7.0 to 0.40 meq/lOOg (Anonymous 1979). Of a very much
wider Ca/Al ratio, however, is the Soiling soil profile in Germany «
described by Ulrich et al. (1971), where exchangeable Ca and Al •
concentrations are respectively 0.2 and 4.7 meq/lOOg, and where
Gb'ttsche's beech studies in the acidification year of 1969 are
plotted to reveal the remarkable correlation between the seasonal •
increase in soluble soil Al concentrations and the dramatic increase w
in fine-root mortality (Ulrich 1980b). Indeed, this correlation and
other studies have encouraged Ulrich (1981) to advance his ecosystem •
hypothesis explaining the widespread "die-back" of fir in Europe. «
Fine roots are killed by high soil Al concentrations or high Al/Ca
ratios with a subsequent invasion of the damaged tree tissues by rot ^
fungi. There is evidence to indicate that increased amounts of •
aluminum can be mobilized in the soil and passed on to water bodies *
(Abrahamsen et al. 1976; Cronan and Schofield 1979). It is not clear
whether the allegedly toxic concentrations present in the •
loess-derived forest soils of central Germany can also be expected to •
arise in the glacial till-derived soils of Scandinavia or
northeastern North America (Tyler 1981).
1
Soil acidification in environments where there is also appreciable
deposition of heavy metals is the second area of concern. Heavy ^
metals arise from various industrial activities, including fuel •
combustion (Hansen and Fisher 1980; Watanabe et al. 1980). The scale *
of emissions and airborne transportation has caused increasing
attention to be directed to the amounts of different elements being I
deposited in remote rural areas. Thus, at the Soiling site in m
central Germany, for an open-site wet deposition of 23.8 kg of
sulphur per hectare per year, there is an accompanying 10 kg of •
nitrogen, 10.4 kg of calcium, 1.9 kg of magnesium and 1.1 kg of •
aluminum (Ulrich 1980b). In south-central Ontario recent comparable
figures are 10 kg for S, 6 kg for N, 5 kg for Ca, and 0.7 kg for Mg -
(Scheider et al. 1979). For the same locality figures for elements •
more commonly understood as "heavy" are 0.46 kg for aluminum, 0.54 kg ™
for iron, 0.095 kg for zinc, 0.132 kg for lead, 0.033 kg for copper
and 0.022 kg for nickel (Jeffries and Snyder 1981). These authors 4|
also point to the much higher deposition rates near smelters where £
cumulative levels of heavy metals in the soils have exercised
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4-63
pronounced toxic effects upon the vegetation (Hutchinson and Whitby
1974, 1976). However, there is danger in extrapolating from such
heavily polluted local situations onto the more diffuse regional
scale without taking into account the different parameters.
Nevertheless, if the Soiling site is taken as exemplifying the more
diffuse rural situation, Heinrichs and Mayer (1977, 1980) found that
the beech and spruce forests act as filters for atmospheric
substances. Some elements (e.g., sulphur, lead, mercury, bismuth and
thallium) are largely accumulated in the upper part of the soil
profile but the complex biogeochemical picture that emerges suggests
that far more needs to be known in other locations on the fate of
deposited metals having potentially significant physiological and
toxicological roles (Andersson et al. 1980a; Bradley et al. 1981;
Smith and Siccama 1981). There is a rapidly expanding literature
focusing on the soil behaviour of heavy metals derived from town
wastes (Leeper 1978) and much of this related to pH-dependent
considerations (Hatton and Pickering 1980; McBride and Blasiak
1979) and metal-organic compound complexes (Bloom et al. 1979;
Marinsky et al. 1980; McBride 1980) should be applicable to the
acidic deposition problem.
The dissolution and mobilization of many other trace metals in soils
is also affected by acidic deposition and decreasing soil pH. Recent
studies in the Adirondack Mountains of New York have determined from
acidic leaching experiments on native bedrock that this process is an
important contributor of Cu, Pb, and Hg in addition to Al (Fuhs
et al. 1981). The trace metals Cu, Pb, Hg, Cd, and Zn were leached
rapidly upon exposure to acid while Al and other major metals were
leached more gradually. Leaching of soils and bedrock by long-term
acidic deposition has resulted in soil impoverishment for metals such as
Mn and Zn in New Zealand (Norton et al. 1981).
Other studies have demonstrated accumulations of trace metals in
soils. Norton et al. (1980) found Pb and chemically similar metals
accumulating in soils while Al and Mn were being leached. Leaching
occurred in the upper soil horizons resulting in potential
impoverishment for shallow rooted plants. Deeper rooted plants, on
the other hand, are subjected to potentially toxic concentrations of
dissolved metals. Tyler (1978) also showed that Pb is not readily
leached from surface soils by acidic deposition inputs. Although the
solubility of this element increases with decreasing pH, most soils
contain sufficient organic matter to tie up the Pb as insoluble
organic - Pb complexes in the soil matrix. Mobility and transport
within the soil horizons and direct atmospheric deposition is
responsible for the accumulation of metals in the soil. For example,
concentrations of Cu and Ni increased in soils with proximity to the
Sudbury, Ontario smelter (Heale 1980). Studies of metal deposition
in the Walker Branch Watershed in Tennessee, found that soils
efficiently retained Pb, Cd, and Cu, and less readily accumulated Cr,
Mn, Zn, and Hg (Andren et al. 1975). McColl (1980), however, found
that the concentrations of Mn, Fe, Cu, and Zn were all greater in
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4-64
soil solutions than in acidic deposition falling in Berkeley,
California.
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In restricted areas, vegetation may be stunted or absent due to
toxicity of metals such as Ni (Foy et al. 1978). In a well-known
study on serpentine-derived soils in Czechoslovakia, Nevmec (1954) M
attributed the failure of pine plantations and various hardwood •
species to excessive levels of Ni , Cr, and Co. Plantation failure
was considerably reduced by fertilization with lime and diabase dust.
Around Sudbury, Canada, Ni and Cu added from atmospheric deposition "W
from smelters are maintained in acidified soils in concentrations •
sufficiently high to be toxic to vegetation (Hutchinson and Whitby
1974, 1976). Thus, any possibility of mobilization of trace metals 1^
through decreasing soil pH by acidic deposition has implications for ^
forest productivity. Accumulations of trace metals from atmospheric
deposition can contribute to this problem.
£
4.5 SENSITIVITY ASSESSMENT
Several sets of sensitivity criteria have been proposed and used to M
define geographical regions most susceptible to acidic deposition
effects (Johnson and Olson in press). Each set of criteria is based ft
upon a different philosophy and is aimed at different target £
organisms or ecosystems (e.g., forests, fish, soil, bedrock, aquatic
ecosystems). Those directed toward aquatic effects have emphasized «
bedrock geology (Hendrey et al. 1980; Norton 1980) or bedrock geology •
and soils in combination (Cowell et al. 1981; Glass et al. 1982; see ™
Section 3.5). Those directed toward terrestrial effects have
emphasized cation exchange capacity and base saturation (Klopatek •
et al. 1980; McFee 1980a,b; Wang and Coote 1981). •
Terrestrial sensitivity has been defined in terms of forest Ife
productivity (Cowell et al. 1981; Table 4-11) and in terms of soil ||
acidification (Wiklander 1973/74, 1980b; Table 4-12). In both cases
effects in the soil body were emphasized. Cowell et al. (1981) ^
regarded low pH soils as the most sensitive based on the assumption ^m
that these already had the smallest reserve of nutrient cations. *
Thus, any additional loss of forest nutrient cations, however small,
would be significant to forest productivity in acid soil systems •
(even though these soils were less sensitive to acidification). This *i
sensitivity assessment concentrated on the upper 25 cm of the soil
profile where, at least in boreal ecosystems, nutrient cycling is •
most efficient. Acid soils are known to actively adsorb SO^", •
hence reduce cation mobilization, and are considered less sensitive
than nonsulphate-adsorbing soils (Johnson and Cole 1977; Singh et al. ^
1980). This contrasts with the sensitivity concept suggested by •
Wiklander (1973/74, 1980b) whereby noncalcareous, moderately acid *
sandy soils (pH 5-6) with low cation exchange capacities are
considered most sensitive. Wiklander (1973/74, 1980b) derived these I
criteria from laboratory studies in which he found that the cation 9
displacing efficiency of H+ was greatly diminished as base
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4-67
saturation and pH decreased. Thus, for a given H+ input, very acid
soils will yield fewer cations and are classed as less sensitive than
moderately acid soils. Moderately acid soils with low cation
exchange capacity (i.e., less buffering by exchange sites) will
experience more rapid pH-change than very acid soils with the same
exchange capacity. This concept of assessing soil acidification
potential, in which the most sensitive soils are those experiencing
the greatest change in their inherent properties, is specifically a
soil sensitivity evaluation. No cause-effect relationships with
vegetative or aquatic systems are specified.
4.5.1 Terrestrial Sensitivity Interpretations
Acidic deposition may cause increases as well as decreases in forest
productivity (Abrahamsen 1980; Cowling and Dochinger 1980). The net
effect on forest growth depends upon a number of site specific
factors such as nutrient status and amount and composition of
atmospheric acid input. In cases where nutrient cations are abundant
and S or N are deficient, moderate inputs of acid may actually
increase forest growth. At the other extreme, acidic deposition in
sufficient amounts may reduce productivity on sites with adequate
N and S but deficient in cations. Other detrimental effects to
forest productivity include changes in soil, microorganisms and Al
toxicity. These effects are increased (Ulrich et al. 1980) with
increased acidification. However, there is insufficient empirical
evidence establishing cause-effect linkages between forest
productivity and acidic deposition. It is not certain which
ecosystem factors are most significant with respect to forest systems
and thus it is not presently possible to map forest productivity
sensitivity at any scale.
Sensitivity assessment for this section, therefore, will concentrate
on soil characteristics and how pH, CEC and sulphate adsorption
properties hypothetically relate to different effects. Terrestrial
ecosystem effects to be considered are: loss of base cations, soil
acidification and Al solubilization. Table 4-13 and Figures 4-5 and
4-6 depict hypothetical sensitivities (Johnson and Olson in press).
For nonsulphate-adsorbing soils (Figure 4-5), it is assumed that each
equivalent of incoming H+ causes the leaching of an equivalent of
some cation (including H+ or Al^+) through the forest soil.
Case 1. For soils with pH >6, H+-base cation exchange is likely
to be nearly 100% efficient (Wiklander 1973/74, 1980b), and thus
soils are very "sensitive" to base cation loss (Figure 4-5a). If the
soil with pH >6 has a high CEC (i.e., a large reserve of
exchangeable cations and hence a large buffering capacity), it will
take a very long time for a given acid input to acidify it. This is
depicted by the width of the CEC box in Figure 4-5a. Thus, such a
soil is thought to have low sensitivity to acidification and Al
mobilization.
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4-68
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>7
6
SOIL 5
4
pH
pH
>7
6
5
4
3-4
> 7
6
SOIL 5
4
3-4
pH
Figure 4-5.
NONSULPHATE-ADSORBING SOILS
(a) CATION EXCHANGE CAPACITY
HIGH
2H+
100
%BASE
SATURATION
804"
CATION
EXCHANGE
BASE
CATIONS
(b) CATION EXCHANGE CAPACITY
HIGH
SO?"
100
% BASE
SATURATION
0
LOW
•4—»>
H+AI3+
CATIONS;
V/////
CATION
EXCHANGE
H+,AI3+
BASE
CATIONS
(c) CATION EXCHANGE CAPACITY
2H+
100
% BASE
SATURATION
H+.AI34"
LOW
//////A///
BASE
CATIONS
CATION
EXCHANGE
H+AI3 +
BASE
CATIONS
V
4-69
INPUT
SOIL
INTERACTIONS
OUTPUT
804" INPUT
SOIL
INTERACTIONS
SO?" OUTPUT
So|" INPUT
SOIL
INTERACTIONS
OUTPUT
Effects on base cation loss, soil acidification and
Al->+ solubilization for nonsulphate-adsorbing soils
having (a) moderate to high pH ( >6), (b) moderate pH
(5-6) and (c) low pH ( <5) (Johnson and Olson in
press).
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4-70
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Case 2. If the soil with pH >6 has a low CEC (area depicted to the
right of the dashed line in the CEC box in Figure 4-5a), it will take
less time to deplete the exchangeable cation reserves and, therefore, •
a low-moderate rating is arbitrarily assigned to acidification and Al f
mobilization in soils to differentiate it from case 1. As in case 1,
H+-cation exchange is nearly 100% complete, so that soils are ^
"sensitive" to base cation loss. •
Case 3. If a soil has pH 5-6 (i.e., a moderate base saturation),
H+-cation exchange will be nearly as complete as in cases 1 and 2 •
while cation reserves (at a given CEC) will be lower (Figure 4-5b). H
For the high CEC case, a moderate rating is assigned to acidification
and Al mobilization in terrestrial ecosystems. As in cases 1 and 2, •
soils are "sensitive" to base cation loss. V
Case 4. In this case, the total reserves of base cations are low _
yet H+-cation exchange is nearly 100% efficient (Figure 4-5b) and I
thus the soil is highly sensitive to base cation loss and ™
acidification. Once base cations are depleted and the soil is
acidified, Al may become mobilized; thus, a moderate rating is I
assigned to soil Al mobilization. •
Case 5. In soils with pH <5 (i.e., low base saturation), H^-base •
cation exchange is less efficient and therefore soils are only 'M
moderately sensitive to base cation loss but have a low sensitivity
to further acidification (Figure 4-5c). Such a soil may be sensitive
to Al mobilization given sufficiently high acid inputs, however. In w
the high CEC case, a moderate sensitivity is assigned to Al P
mobilization in soils.
1
Case b. In this case, the soil is acid and has low CEC (Figure
4-5c), making it only moderately sensitive to base cation loss but
highly sensitive to Al mobilization. These soils have only a low •
sensitivity to further acidification. •
In sulphate-adsorbing soils, leaching of H"*", Al^+, and base
cations is inhibited for the reasons described previously. The •
degree of sulphate adsorption is dependent to some extent on pH 9
(Harward and Reisenauer 1966) as well as on inherent soil properties
such as organic matter and Fe- and Al-oxide content. Sulphate is 1|
more strongly adsorbed in most acid soils. Soils having a pH >6 |
would not be expected to exhibit high sulphate adsorption properties
(cases 7 and 8; Table 4-13). High Fe- and Al-sesquioxide content ^
required for sulphate adsorption would not be characteristic of these •
soils. Also, any soils having high CEC (cases 9 and 11; Table 4-13) *
would not be expected to adsorb sulphate if the exchange capacity was
controlled primarily by organic matter. High organic matter content •
tends to block the adsorption potential of Fe and Al oxides (Johnson •
and Henderson 1979; Singh et al. 1980).
Cases 9 to 12 are analogous to cases 3 to 6, respectively, with •
regard to pH and CEC. Due to the sulphate adsorption capacity of
1
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4-71
soils in cases 9 to 12, however, base cation leaching is lower than
in analogous soils in cases 3 to 6 (Figure 4-6). However, sulphate
adsorption prevents H transport as well as base cation leaching
from soils in cases 9 to 12, and the displacing efficiency of H
for base cations is less of a factor in soil acidification. Soils in
cases 9 to 12 are rated slightly more sensitive to Al solubilization
than their counterparts 3 to 6.
The preceding discussion as well as Table 4-13 and Figures 4-5 and
4-6 by no means represent a rigid set of criteria for site sensiti-
vity to acidic deposition. They merely represent a hypothetical
construct based upon a combination of known and previously employed
sensitivity criteria. These criteria have limitations and are by no
means complete. For instance, while considering the sensitivity of
terrestrial and aquatic ecosystems to acidic deposition effects, it is
important to realize that acids are produced naturally within these
systems (Rosenqvist 1978). The effects of atmospheric acid inputs
must be viewed as an addition to natural, ongoing acidification and
leaching processes within soils due to carbonic acid formation,
organic acid formation, free cation uptake, and a variety of other
processes (Andersson et al. 1980b; Sollins et al. 1980; Ulrich 1980a).
Unfortunately, the data base for including natural acid formation
criteria into regional sensitivity assessments is extremely limited.
Thus, previous sensitivity rating schemes, by default, assume that
atmospheric inputs add significantly to internal acid production, an
assumption that is by no means universally accepted (Rosenqvist
1978).
It is also important to distinguish between acidification and
elemental leaching when considering the role of natural acid
formation. Carbonic acid is a major leaching agent in some forest
soils, yet it does not produce low pH (i.e., <5.0) solutions under
normal conditions (Johnson et al. 1977). Organic acids may contribute
substantially to elemental leaching in forest soils undergoing
podzolization (Johnson et al. 1977), and they can produce low pH
(i.e., <5.0) in unpolluted natural waters as well (Johnson 1981).
Also, because leaching is only one of several processes that affect
soil acidity (other major factors being humus build-up, plant cation
uptake, and mineral weathering; Ulrich 1980a), the relative
contribution of atmospheric acidic deposition to elemental leaching
may be quite different from the relative contribution of atmospheric
deposition to soil acidification.
4.5.2 Terrestrial Mapping for Eastern North America
The lack of empirical data identifying cause-effect linkages with
respect to impacts of acidic deposition on forest and agricultural
systems makes sensitivity mapping difficult. One must identify a
target process (i.e., soil acidification, Al solubilization, cation
loss, etc.) and then make specific assumptions regarding ecosystem
interactions which result in a significant impact. The mapping
-------
4-72
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>7
6
PH 5
4
3-4
>7
6
pH 5
4
3-4
SULPHATE-ADSORBING SOILS
(a) CATION EXCHANGE CAPACITY
,Fe , Al OXIDES
2H +
r^y
100
%BASE
SATURATIONS
SO?"
ANION
ADSORPTION
CATION
EXCHANGE
INPUT
SOIL
INTERACTIONS
H+,AI3+
BASE
CATIONS
S04~
OUTPUT
(b) CATION EXCHANGE CAPACITY
•Fe , Al OXIDES
HIGH
100
%BASE
SATURATION
0
H+,
LOW
r*
(
i
i
AI3+j
•1
// ///A/}LS
2-
M/
ANION
'ADSORPTION
CATION
EXCHANGE
INPUT
SOIL
INTERACTIONS
BASE CATION"
H+,AI3 +
BASE
CATIONS
SO
2-
OUTPUT
Figure 4-6.
Effects on base cation loss, soil acidification and
Oj_
Al solubilization for sulphate-adsorbing soils
having (a) moderate pH (5-6) and (b) low pH (<5)
(Johnson and Olson in press).
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4-73
framework and thus the maps themselves could vary substantially
depending on these assumptions and on which processes were targeted
(Table 4-13). For this report mapping is concentrated on terrain
characteristics, especially soil chemistry (or its surrogate), of
eastern North America rather than sensitivity per se. It is hoped
these maps will provide a useful basis for comparison and for
eventual interpretations of terrestrial sensitivities once actual
impacts are documented.
Figures 4-7 and 3-9 (in map folio) show terrain characteristics in
the Eastern United States and Eastern Canada, respectively. Although
an attempt was made to compile compatible and comparable maps,
differences in mapping criteria and data quality exist between and
within the two maps. Specific mapping factors and data sources are
listed in Table 4-14. The data availability in some parts of eastern
Canada necessitated map production at 1:1,000,000 to ensure adequate
representation. The U.S. map, produced from the Geoecology Data Base
(Olsen et al. 1980) is reproduced here at 1:5,000,000.
Terrain characteristics for the United States are mapped as soil
chemical classes based on pH and CEC. These classes can be compared
back to Table 4-13 to interpret relative sensitivities to acidic
deposition for Al solubilization, cation loss and soil acidification.
The discussion which follows deals primarily with potential for soil
acidification based on the Wiklander (1973/74) concept of
sensitivity. The limit of Wisconsin glaciation is shown on the map
as a basis for comparison between younger soils of the north and
northeast and the older, more deeply weathered soils characteristic
of the south and southeast.
Bedrock geology has been included in terrain classes for the
Canadian mapping because soils in Canada, especially throughout the
Precambrian Shield, tend to be thin and discontinuous. In these
areas the bedrock forms a major substrate for forest systems and
represents the only "store" of nutrient cations. Because of a lack
of soil chemical data for soils outside limited agricultural areas in
Canada, soil texture and depth to carbonate data have been
substituted (Table 4-14). These are the only surrogates available
for soil chemistry at the scale of compilation (1:1,000,000). It is
not possible, therefore, to relate the terrain classes on the
Canadian map directly to the soil property classes shown in
Table 4-13. However, some indirect comparisons can be made on the
bases of soil order.
4.5.2.1 Eastern United States
The map showing various classes of soil characteristics covering the
eastern 37 states (Figure 4-7) was produced at Oak Ridge National
Laboratory (Olson et al. 1982). The analysis utilized various
national resource inventories to map soil classes based on pH, CEC,
Histosols and land use (Table 4-14). County-level data from the
Geoecology Data Base (Olson et al. 1980) were used in the analysis to
provide a regional perspective and understanding of soil
-------
4-74
characteristics which can be evaluated in terms of the potential for
acidic deposition impacts on terrestrial ecosystems (Table 4-13). As
more detailed data or new studies are completed, the resolution or
interpretation of the map may need to be revised.
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Initially counties that were predominantly ( > 50%) urban or ,»
agricultural were excluded from the analysis. Management and land •
use practices (e.g., liming, fertilizing) in these areas would tend
to determine current soil nutrient/pH conditions. The 1977 National
Resource Inventory (USDA 1978) was used to define land in urban •
built-up areas and transportation corridors. The 1978 Census of •
Agriculture (USDC 1979) provided data on cropland. This resulted in
1,648 of the 2,660 counties in the east being included in the
analysis as containing predominantly forest, range, or pasture.
M
•
Moderately acid soil (pH 5-6) and low CEC were used to identify soils _
which would be most sensitive to acidification (Wiklander 1973/74, •
1980b). Very acid soils will yield fewer cations and thus, are *
classed as less sensitive than moderately acid soils (Wiklander
1973/74). Moderately acid soils with low CEC (i.e., less buffering I
by exchange sites) will experience more rapid pH change than very m
acid soils with the same CEC (Table 4-13). McFee (1980a,b) used
CEC only as a first reasonable approximation to site sensitivity. He •
used a CEC value of 6.2 meq/lOOg to identify soils that would change •
most quickly under the influence of acidic deposition. This
criterion was used to distinguish between "low CEC soils" and "high ^
CEC soils" (Table 4-9). As noted earlier, however, the interpre- •
tation of soils sensitive to acidic deposition does not account for •
the relative significance of this deposition compared to internal
acid generation. It is not presently possible to classify soils as •
to internal acid generation especially in a regional-level analysis. £
Counties covered by 50% or more of soil types with a surface pH of ^
greater than 5.5 and CEC less than 6.2 were classified as having a •
high potential to undergo acidification from acidic deposition. In
addition, two other classes were defined which may undergo
acidification but at a slower rate with similar levels of acidic •
deposition. Soils with a pH greater than 5.0 and CEC less than 6.2 9
constituted class 2. Class 3 included soils with pH greater than 5.0
and CEC less than 9.0. •
Chemical and physical soil characteristics employed in the analysis
represent average values for the A horizon (upper 20-25 cm) for the ^
82 great soil groups occurring in the eastern United States. These •
values were obtained from published literature (Klopatek et al. ™
1980). The great soil groups were combined to estimate values for
the 195 soil mapping units that are mapped (USGS 1970) in the east. fl
Although the exact proportions of great soil groups within map units P
are not readily available, the dominant great soil group was given a
weighting factor of 0.66 to calculate average map unit values. M
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4-75
Proportions of soil mapping units within counties were estimated from
the 1:7,500,000 scale soil map of the United States (USGS 1970).
Sulphate adsorption capacity was also considered in defining
sensitive soils. Soils with sulphate adsorption capacity prevent the
transport of cations by H2S04 (Johnson and Cole 1977) and can
increase soil exchange capacity (Wiklander 1980a) thus reducing the
sensitivity of such soils to acidification. Because of the relative
lack of empirical information on sulphate adsorption capacity, only
Ultisols are considered as sulphate-adsorbing soils (Johnson et al.
1980). Even the data on sulphate adsorption capacity of Ultisols are
limited in geographic extent. Ultisols have lower pH and higher CEC
values than given above for defining soils sensitive to acidic
deposition. Therefore, sulphate adsorption does not appear on Figure
4-7 as part of the classification. Soils other than Ultisols have
varying capacities to adsorb sulphate and this factor may be
important in mediating the acidification of a soil that would
otherwise appear sensitive.
Counties were used both to integrate the various factors and to
display the results. Although counties are generally uniform in size
in the eastern United States, some of the larger counties occur along
the U.S./Canada border in Maine and Minnesota. All the factors
utilized a 50% criteria to classify counties. Therefore, significant
areas can exist within counties that differ from the final designated
classification. Thus Figure 4-7 displays the broad regional patterns
but evaluation of an individual county requires more detailed
analysis to determine the extent and coincidence of the various
factors within that county.
Six classes were used to characterize soils (Table 4-15), with
agricultural/urban areas shown as blank on the map (Figure 4-7).
Classes 1 to 3 are specifically related to the increasing potential
for soils to undergo acidification with acidic deposition. The other
three classes were included to provide additional information on
soils which could be used with alternate hypothesis of soil
sensitivity or in better understanding acid inputs from soils to
aquatic systems (Section 3.5). Class 4 includes low pH soils
(pH <5.0) and Class 5 includes high pH soils (pH > 5.0) that also
have a high CEC (CEC >9.0). Class 4 soils are those most likely to
experience Al mobilization and have the potential to transfer both
H+ or Al3+ ions to aquatic systems, given sufficient inputs of
acid. Section 3.5 with Figure 3-10 provides additional discussion on
the potential transfer of acid to aquatic systems.
Class 6 includes areas dominated by Histosols (peat soils). Although
these organic soils may not be sensitive to further acidification, it
is important to recognize them in the overall assessment of acidic
deposition impacts. Class 6 is most informative when compared to
Figure 3-10 in Section 3.5 relative to acid inputs to aquatic
systems. These and similar Canadian peatland areas naturally
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4-76
I
TABLE 4-14. TERRESTRIAL FACTORS AND ASSOCIATED DATA BASES UTILIZED FOR
TERRAIN CHARACTERISTICS MAPPING IN EASTERN CANADA AND THE
EASTERN UNITED STATES
Terrestrial Factors/Surrogates
EASTERN UNITED STATES
1) Soil Chemistry
i) Mean soil order pH
( in distilled water)
ii) Mean soil order CEC
iii) Histisols (Organic soils)
2) Limit of Wisconsin Glaciation.
EASTERN CANADA
1) Soil Chemistry
Surrogates: i) Texture (sand, loam
or clay) - Northern
Ontario, Quebec,
the Maritimes and
Newfoundland/Labrador .
ii) Depth to Carbonate
(high, low or no
iii) Glacial Landforms -
northwestern Ontario..
iv) Organic Soils (>50%
of mapping unit) ......
2) Soil Depth - very shallow (approx. <25 cm),.
- shallow and deep C25 cm)
3) Bedrock Geology — lithology. ................
a All U.S. data sources listed have been compiled
Data Base (Olson et al. 1980).
Data Sources3
..Soil Map (USGS 1970)
..Soil Map (USGS 1970)
..Soil Map (USGS 1970)
..USGS 1970
..1977 National Resource
Inventory (USDA 1978)
. . 1978 Census of
Agriculture (USDC 1979)
. .Ecodistrict Data Base
(Environment Canada
1981a,b,c)
. .Ontario Land Inventory
(OMNR 1977)
..Pala and Boissonneau 1979
. .Ecodistricts (Environment
Canada 1981a,b,c)
..Ecodistricts (Environment
Canada 1981a,b,c) and
Ontario Land Inventory
(OMNR 1977)
..Shilts et al. 1981
. .Ecodistricts (Environment
Canada 1981a,b,c) and
Ontario Land Inventory
(OMNR 1977)
within the Geoecology
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TABLE 4-15.
4-77
SOIL CHEMICAL CLASSES AND AREAS DOMINATED BY HISTOSOLS IN
THE EASTERN UNITED STATES AS MAPPED IN FIGURE 4-7
Class
Acidification
Potential
No. of
Counties
Characteristics
1 High 16
2 Moderate 19
3 Moderate-Low 45
4 Low 849
5 Low 700
6 Low 13
Moderate pH (>5.5), lowest CEC (<6.2)
Moderate pH (>5.5), low CEC (<9.0)
Moderate pH (>5.0), low CEC (<9.0)
Low pH (>5.0)
Moderate pH (>5.0), high CEC (>9.0)
Histosols
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4-78
Category I: Organic Soils (Histosols)
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contribute acids of varying strength to lakes and streams due to
high levels of organic acids.
4.5.2.2 Eastern Canada
The map of eastern Canada is the same as that used in Chapter 3 ||g
(Figure 3-9). The reader is referred to Section 3.5.1.1 for detail
regarding map compilation. ^
This map shows 62 classes of terrain characteristics based on ™
combinations of soil depth, soil texture (or depth to carbonate) and
bedrock sensitivity. In the previous chapter it was used to define fl
areas of varying potential to reduce the acidity of incoming f
precipitation prior to entering surface waters. In this section the
62 map classes are recombined in order to emphasize soil •
characteristics. The classes have been grouped according to five •
soil categories: (1) organic soils (I); (2) barren areas (>75%
bedrock outcropping; II); (3) sandy soils (III); (4) loamy soils
(IV); and (5) clayey soils (V; Table 4-16). These are correlated to m
soil orders of the Canadian System of Soil Classification. They are •
also subdivided on the basis of underlying bedrock sensitivities. It
is possible to break down these divisions further, such as by soil
depth (25 cm - 1 m and >1 m), or even to recombine classes if some
other characteristic is preferred as a basis for discrimination.
However, the grouping suggested in Table 4-16 provides a framework «
for summarizing soils of eastern Canada and discussing some aspects •
of terrestrial sensitivity.
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In Table 4-16, organic soils overlying carbonate bedrock are
identified separately (IA) from those overlying other rock types A
(IB). Category IB map units occur over wide areas of Ontario, Quebec fl
and Labrador and in small pockets throughout eastern Canada. Organic
soils overlying carbonate strata is most common in the Hudson Bay ^
Lowland of northern Ontario and northwestern Quebec. •
The presence of limestones and dolomites beneath peat is significant
because local hydrological conditions influence peat development and A
peatland chemistry. More minerotrophic types of peatland ecosystems 9
occur where groundwater comes in contact with the substrate (Cowell
et al. 1979; Sjors 1963). Consequently large portions of IA organic •
soils have "fen" and "swamp" type ecosystems which may have a •
groundwater pH well in excess of 5. This could be a significant
consideration with respect to the impact of acidic deposition on such
terrain. This is also true for IB peatlands (and other wetlands) •
occurring on clay (northeastern Ontario and southern Ontario). ™
Peatlands occurring on the Canadian Shield tend to be less
minerotrophic but local soil/groundwater conditions need to be
evaluated.
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Category II: Barren Areas
These are map units dominated (^75%) by exposed bedrock. Three
classes of barren terrain have been mapped on the basis of bedrock
lithology identified in Table 4-11 according to "sensitivity".
Levels of sensitivity relate the potential or capacity of rock types
to reduce acidity of atmospheric deposition as defined by Lucas and
Cowell (1982). Most of these classes in eastern Canada lie above the
treeline and southern limit of continuous permafrost.
Category III: Sand or No Lime Soils
According to the Soil Map of Canada (Clayton et al. 1977) these areas
(IIIA to ITIC; Table 4-16) are primarily Humo-Ferric Podzols,
"Rockland" ( j^60% exposed bedrock) and Dystric Brunisols. These
soil types have acidic surface horizons (pH <5.5, dominantly <5)
and correlate most closely with cases 5 and 6 in Table 4-13. They
are thus considered to have a low sensitivity to acidification, a
moderate sensitivity to base cation loss and a moderate to high
sensitivity with respect to Al solubilization. Soils mapped as Mlq,
Mir, M4a and M4b (IIIB), and L2d and L3 (IIIC) in Figure 3-9 are
probably the most sensitive because they are the shallowest.
Category IIIA soils however overlie calcareous bedrock.
Boreal and northern temperate podzols are characterized by the
accumulation of organic matter and Fe-and Al-sesquioxides (Stobbe
1968). Although high Fe and Al content are properties known to
enhance sulphate adsorption (Johnson and Cole 1977), high organic
matter tends to block the adsorption process (Johnson and Henderson
1979). Low pH, high CEC podzols in eastern Canada probably do not
adsorb sulphate significantly (Case 11; Table 4-13) because their
CEC is controlled primarily by organic matter. At this time however,
there is very little empirical evidence regarding the sulphate
adsorption capacity of Canadian soils.
Category IV: Loam or Low Lime Soils
Categories IVA to IVC are represented by loam/low lime soils of
varying depth and overlie different bedrock types. It is not certain
how these relate to the soil chemical classes identified in Table
4-13. They are mapped by Clayton et al. (1977) as primarily Podzolic
and Brunisolic (and as Rockland). Based on the interpretation of
texture and depth to carbonate as surrogates for soil chemistry,
these classes are considered to exhibit the properties of Cases 3 and
4 in Table 4-13 (moderate pH).
Category V: Clay or High Lime Soils
These soils (VA to VC) are interpreted as having low to moderate
sensitivity with respect to soil acidification and Al
solubilization. However, sensitivity to base cation loss is high.
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4-82
incidence of common diseases and insect infestations is likely
to be affected by acidic deposition and ozone.
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These soils are primarily Gray Luvisols and Gleysols (clay-rich
and/or under periodic or seasonal flooding) according to the Soil Map
of Canada (Clayton et al. 1977). •
Map units in Figure 3-9, as noted earlier, are based on Ecodistrict
and Ontario Land Inventory polygons. The best spatial resolution is
in Ontario, south of 52 °N latitude, where Ontario Land Inventory
(OMNR 1977) units were used. Because soil and bedrock
characteristics are identified on the basis of dominance, subdominant •
characteristics are not shown on the map. There remains a need for •
much improved soil chemistry data from nonagricultural areas in
eastern Canada.
4.6 RESEARCH NEEDS
The following list does not confer an order of priority; rather, the f
ordering reflects the general progression of the foregoing chapter
from INTRODUCTION through SENSITIVITY ASSESSMENT. m
I
1. Improve resolution (spatial and temporal) of current wet and dry "
deposition patterns in both the United States and Canada.
2. Improve projection of wet and dry deposition within designated
areas of United States and Canada.
3. Improve information on capture and fate of dry S and N within Q
principal terrestrial ecosystem types.
4. Determine tree and crop species exposed to greatest risk of •
reduction in productivity by acidic deposition. Determine plant ~
characteristics associated with susceptibility/tolerance to 03
and acidic deposition. •
5. Determine quantitative relationships between dose-response
acidic deposition and productivity of trees and crops. •
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6. Determine extent to which dose-response relationships are
altered by presence of 03, deposition of particulates, soil _
nutrient and moisture supplies, and pattern and timing of W
precipitation events with respect to stages of plant develop- ™
ment. Identify stages of vulnerability of agricultural crops
and/or forest vegetation, particularly to episodic wet and/or ft
dry deposition. |
7. Determine degree to which uptake of metals, particularly tm
aluminum, is increased by exposure to acidic deposition. •
8. Determine the interaction of acid stress with other abiotic and
with biotic stresses on terrestrial plants. Determine whether •
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4-83
9. Develop a standardized biological indicator of acidic
deposition, having known relationships to changes in
productivity of trees and crops.
10. Identify beneficial as well as injurious effects of various
components of acid pollution, with particular reference to rate
relationships for principal terrestrial ecosystem types.
Determine the effects of H4", SO^", and NC>3~, separately and
combined, on forest nutrient status. (This is a problem of
quantifying benefits of SO^" and N03~ inputs vs.
detriment of H+ deposition and net overall effect on forest
ecosystems at current and projected input levels.)
11. Based on actual field observations, quantify natural H+ ion
production and consumption rates for the principal terrestrial
ecosystem types, and the clear distinction of anthropogenic and
natural H+ ion production. Obtain more information on natural
internal acid production and leaching for a variety of forest
ecosystems. (This must be used in a full, comprehensive
analysis of acidic deposition effects on soil leaching of metal
cations and transfer to aquatic ecosystems.)
12. Determine sensitivity of aquatic and terrestrial components of
headwater pond and lake ecosystems to acid loadings.
13. Improve information, based on actual field observations on a
representative range of soil types, on impact of acidic
deposition on sensitive biochemical and/or chemical processes,
and generally identify soil types sensitive to various
pollutants and pollutant combinations.
14. Determine major factors affecting soil SO^" adsorption
capacity, and how they vary among soil orders and/or major soil
types.
15. Based on actual field observation on a representative range of
mainly natural soils, improve information on impact of acidic
deposition on soil biota, soil mineralogy and soil organic
matter.
16. Improve understanding of relationships between forest
productivity and acid sensitive properties of soils.
17. Consider the long-term site impoverishment potential of
continued acidic deposition, in the light of trends in forest
management toward more rapid growth, shorter rotations and
full-tree, or even whole-tree, harvesting.
18. Improve system of mapping terrestrial sensitivity, hopefully
incorporating existent data bases, to allow further
identification of key sensitive areas.
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4-84
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19. Improve soils information base, including also such factors as
depth to carbonate and sulphate adsorption capacity,
particularly in remote areas. •
20. Extended or "whole" ecosystem field studies of biological
response to acidic deposition to understand the complex trophic •
interactions amongst the organisms and the resulting community fj
changes, particularly those affecting wildlife and the top
carnivores in the food chain. ^
•
21. Develop mitigative measures for correcting acidity impacts. ™
22. Determine the "threshold" or critical dosage of 03 required to •
produce injury and/or suppression of growth and yield under a m
variety of field conditions.
23. Determine interactive response involving 03 and chemical £
additives (i.e., insecticides, fungicides, nematocides, growth
regulators). When responses occur, identify physical and ^
chemical factors involved in the interactions. •
24. Determine the diurnal pattern of 03 occurrence in the major
agricultural and forested regions as a guide for field Aj
fumigation studies and as a guide for calculation of realistic £
dosages.
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4.7 CONCLUSIONS
1. Field and laboratory studies with 03 that indicate reductions m
in yield may occur for various tree species and such crops as »
beans, tobacco, potatoes, onions, radishes, grapes, soybeans and
sweet corn. During the growth season frequent exposures to 03 A
concentrations in excess of 0.1 ppm have produced up to 20% yield ||
losses for susceptible species.
2. Although simulated rainfall experiments have produced some direct V
effects on plants exposed to higher than normal H+ concentra- *
tions, direct effects have not been documented conclusively in
the field for vegetation exposed to ambient precipitation. •
3. Experiments with simulated acidic deposition and 03 have
demonstrated greater plant growth reduction from the two together A
than would be expected from the results of their individual •
effects.
4. Individual precipitation events which occur during critical H
growth stages (e.g., during flowering or pollination) offer "
amplified potential for damage to agricultural crops.
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4-85
5. Direct effects of acidic deposition on soils have been shown to
increase 8042- movement and increase the rate of nutrient
cation denudation. However, some soils exhibit a substantial
capacity to adsorb SO^j and resist nutrient cation
leaching.
6. The terrestrial system's influence on the acid component of
atmospheric deposition has important implications for the aquatic
ecosystem.
7. Multifactor data bases have been employed, to develop maps of
eastern North America which depict the sensitivity of various
areas (down to the county level for the U.S. and Ecodistricts in
Canada) to impacts from acidic deposition.
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4-86
4.8 REFERENCES
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Abrahamsen, G. 1980. Acid precipitation, plant nutrients and forest •
group. In Proc. Int. Conf. Ecological Impact of Acid f
Precipitation, eds. D. Drablos and A. Tollan, pp. 58-63.
SNSF-Project, Sandefjord, Norway, 1980. M
Abrahamsen, G.; Horntvedt, R.; and Tveite, B. 1976. Impacts of acid
precipitation on coniferous forest ecosystems. In Proc. First
Int. Symp. Acid Precipitation and the Forest Ecosystem, eds. M
L.S. Dochinger and T.S. Seliga, pp. 991-1009. USDA Forest »
Service Gen. Tech. Report NE-23, Columbus, OH., 1976.
. 1977. Impacts of acid precipitation on coniferous |
forest ecosystems. Water, Air, Soil Pollut. 8:57-73.
Abrahamsen, G., and Stuanes, A.D. 1980. Effects of simulated rain •
on the effluent from lysimeters with acid, shallow soil, rich in
organic matter. In Proc. Int. Conf. Ecological Impact of Acid
Precipitation, eds. D. Drablos and A. Tollan, pp. 152-153. •
SNSF-Project, Sandefjord, Norway, 1980. •
Alexander, M. 1980. Effects of acidity on microorganisms and •
microbial processes in soil. In Effects of acid precipitation f
on terrestrial ecosystems, eds. T.C. Hutchinson and M. Havas,
pp. 363-374. New York: Plenum Press. ^
Allaway, W.H. 1970. Sulphur-selenium relationships in soils and ™
plants. Sulphur Inst. J. 6(3):3-5.
_
forage and pastures: selenium in forages as related to the
geographic distribution of muscular dystrophy in livestock. M
J. Anim. Sci. 23:271-277. g
Andren, A.W.; Lindberg, S.E.; and Bate, L.C. 1975. Atmospheric
input and geochemical cycling of selected trace elements in •
Walker Branch Watershed. Environmental Sciences Div. Pub. No. •
728, Oak Ridge National Laboratory, Oak Ridge, TN.
Andersson, A.M.; Johnson, A.H.; and Siccama, T.G. 1980a. Levels of ^
lead, copper and zinc in the forest floor in the northeastern
United States. J. Environ. Qual. 9:293-296. M
Andersson, A.M., and Nilsson, K.O. 1974. Influence of lime and soil ~
pH on K availability to plants. Ambio 3(5):198-200.
Andersson, F.; Fagerstrom, T.; and Nilsson, I. 1980b. Forest V
ecosystem responses to acid deposition - hydrogen ion budget and
nitrogen/tree growth model approaches. In Effects of acid •
precipitation on terrestrial ecosystems, eds. T.C. Hutchinson ^
and M. Havas, pp. 319-334. New York: Plenum Press.
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SECTION 5
HEALTH AND VISIBILITY
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5-1
SECTION 5
HEALTH AND VISIBILITY
5.1 HEALTH
A complete assessment of the health implications of U.S./Canadian
transboundary air pollution would encompass the full range of current
pollution concerns, including photochemical oxidants, sulphur and
nitrogen oxides, particulate matter, and associated toxic substances.
Although future phases may address these air quality concerns more
completely, this report will focus on potentially indirect health
effects associated with the transboundary deposition of acidifying
substances, with only a brief summary of information on direct
inhalation of the above mentioned pollutants.
Available information gives little cause for concern over direct
health effects from acidic deposition. The pH of acidic deposition
is generally well within the range normally tolerated by the skin and
gastrointestinal tract. Although high levels of SC^, NOo, and
acidic aerosols are reported in urban areas, no studies have been
found which suggest adverse effects from dry deposition on the skin.
Evidence does suggest, however, that inhalation of high levels of
such substances may produce respiratory and other internal disease
(NAS 1977a; USEPA 1980), and one early epidemiological study (Gorham
1958) even reported an inverse statistical association between
bronchitis mortality and the pH of winter precipitation in Great
Britain. In this case precipitation acidity was probably an index of
acid precursor air quality, since a plausible mechanism for causality
does not exist.
Evidence for the following potentially indirect health effects
associated with acidic deposition is discussed below:
(1) contamination of edible fish by toxic materials, principally
mercury; (2) leaching and corrosion of watersheds and storage and
distribution systems, leading to elevated levels of toxic elements;
and (3) prolonged direct contact with acidified water in recreation
settings. In addition, a brief assessment is given on direct
inhalation of common pollutant classes (i.e., photochemical oxidants,
acidic aerosols and sulphur and nitrogen oxides) that can be
associated with acidic deposition.
5.1.1 Contamination of Edible Fish
Some evidence suggests that acidic deposition may alter the
biogeochemical cycle of metals, including mercury (Brosset and
Svedung 1977; Jensen and Jernelov 1972; Schindler 1980; Tomlinson
1978). Poorly buffered waters in areas remote from any point source
of discharge of mercury have been found to contain fish with elevated
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levels of mercury. Scheider et al. (1979) found that the mercury
content of walleye from 21 Ontario lakes was significantly higher for _
lakes where alkalinity was less than 15 mg/L (as CaCOo), as opposed •
to lakes with higher alkalinities. According to Tomlinson et al.
(1979), fish from poorly buffered lakes and rivers in Quebec, New
Brunswick, Minnesota, New York, and Maine also contain elevated I
mercury. I
The mechanisms by which acidic deposition might increase fish mercury •
content are not known, but most likely involve both biological and I
chemical processes. The principal forms of mercury of interest are
elemental (Hg°), dimethyl mercury ((CHo^Hg), mercuric mercury _
(Hg ), and monomethyl mercury (CH-jHg"*). Jensen and Jernelov I
(1972), and several other investigators, have shown that inorganic ™
mercury can be methylated in both aquatic and terrestrial ecosystems.
One hypothesis that attempts to explain the relationship between pH •
and mercury, holds that monomethyl mercury formation is at low pH |
( <7), while dimethyl mercury forms at higher pH ( > 7) (Jensen and
Jernelov 1972; Tomlinson et al. 1979). Dimethyl mercury has a high «
vapour pressure, is relatively insoluble, and is thus largely •
released to the atmosphere. Methyl mercury uptake by fish in lakes *
having higher pH regimes would thus be minimized. According to this
hypothesis, lakes with lower pH produce proportionately larger I
amounts of monomethyl mercury, which is efficiently taken up by I
biota. The reduced availability of young fish containing low mercury
levels, and increased foraging activity by larger predator fish, both ••
characteristic of acidified lakes, would then increase the •
bioaccumulation of methyl mercury in larger fish. Recent
experiments, however, found a very poor correlation between mono- and —
dimethyl mercury versus pH, calling this hypothesized mechanism into •
question. ™
The processes leading to increased mercury burdens in fish are likely •
to be more complex, including considerations like the complexity of p
the food chain, redox conditions, inorganic and organic sequestering
agents, watershed to lake area ratio (Suns et al. 1980), as well as *m
the rate of atmospheric mercury input. •
Although natural sources appear to contribute the major portion of
atmospheric mercury (NRC 1978), emissions from coal combustion can be I
of significance on a regional scale. Lindberg (1980) collected air •
samples from a plume of a major coal-fired generating station and
found that the mercury emitted was predominantly in the elemental •
vapour phase, with very little conversion to particles as the J|
distance from the source increased. Due to the nature of the Hg, his
findings support the theory that the majority of Hg emitted during _
coal combustion is deposited regionally rather than locally. Since I
the Hg is in the vapour phase, the major route of removal is thought •
to be via precipitation scavenging, which should theoretically
increase in efficiency as the pH of the precipitation falls. Using •
calculations from Brosset and Svedung (1977), Tomlinson (1978), and |
Brouzes et al. (1977), it is hypothesized that acid-containing
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5-3
clouds and rains should effectively remove methyl mercury from the
atmosphere. The surface of acidified lakes should also be an
effective sink for the dry deposition of methyl mercury. Once
removed, methyl mercury would then be more likely to stay in solution
in acidified waters. This scenario is shown in Figure 5-1, which
illustrates the distribution of mercury in waters of three different
acidities. Again, recent evidence suggests that the process is much
more complicated (Barton and Johnson 1980). These authors could
detect no dimethyl mercury in air from measurements taken at several
locations in Ontario. Although a clear association appears to exist
between mercury in fish and acidity of lakes, the mechanism by which
this phenomenon might be explained remains obscure. Although the
extent to which acidic deposition may have contributed to
mobilization or retention of mercury in fish is speculative, fish
that are harvested from these lakes present a potential health hazard
to humans.
Presently the mercury content of fish tested in the affected areas is
usually less than the U.S. FDA recommended levels of 0.50 yg/g. If
the situation is not changed, it would be prudent to assume that the
mercury content of fish will continue to rise as lake pH drops
(Figure 5-2).
Another important factor is that the bioaccumulation of mercury is
related to the species' trophic level. The larger pisciverous fish
are known to have greater concentrations of mercury in the tissues
than the planktivors (Philips et al. 1980). These fish are also the
most prized sport fish, and make up the majority of the yearly catch
eaten. Little research has been directed at mammals inhabiting areas
of elevated mercury levels. One study by Wren et al. (1980) suggests
that terrestrial species have a demethylating process, which can
reduce the amount of toxic organic mercury in their bodies.
With respect to human health, elevated levels of mercury can lead to
serious disorders. The severity of these ailments is usually related
to the exposure level to mercury. This type of disorder has been
reviewed in the past by many authors including Chang (1977).
Generally the blood-Hg level for threshold effects lies somewhere
between 100 and 200 ng/mL in a normal adult, but the maximum
recommended blood Hg level for pregnant females is 20 ng/mL (NRCC
1979).
Epidemiological studies have been completed in Canada which
investigated the health of populations, especially natives, that were
exposed to increased concentrations of Hg in food and had as a
result, elevated blood and hair Hg levels (Rudy 1980). For example,
Rudy (1980) documented some mild neurological abnormalities in adult
Cree men and an association between reflexes in young Cree boys and
the concentration of methyl mercury in their mothers' hair during
pregnancy. Due to the lack of accurate exposure modelling and many
shortcomings, it would be premature to regard this work as an example
of a long-range transport of air pollution associated problem.
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5-4
(CH3)2Hg
.1
Sphagnum
LAKE pH
Mercury
Concentration
In Fish
<5.0
No Fish
3.5- 6.5
7.5-8.0
CaCO-:
0.5 -5.0ppm(mg/l) 0.1 —1.0 ppm(mg/D
Figure 5-1. Varying effects of lake pH on the distribution of
mercury in ecosystems (Tomlinson 1978).
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200
180
160
O)
O)
140
DC
I-
Z
LU
O
120
O 100
O
O 80
DC
LU
^.
60
40
20
4.0
5.0
5.5
6.0
PH
5-5
A = 0.63
p<0.05
6.5
7.0 7.5
Figure 5-2. Mercury in yearling yellow perch and epilimnetic pH
relationships (Suns et al. 1980).
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5-6
However, an appreciation of the potential risk of long-term exposure
to elevated levels of Hg in food should be maintained and careful
monitoring of the situation should continue to avoid any further
deterioration in health.
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5.1.2 Contamination of Drinking Water •
Acidic deposition can increase the concentration of toxic metals in
drinking water by: (1) increasing the deposition of metal in soluble V
forms (e.g., mercury); (2) leaching of metals from the watershed and 9
from sediments; and (3) acid corrosion of materials used in
reservoirs and drinking water distribution systems. •
Again, no clear evidence of health effects arising from the
consumption of drinking water contaminated with metals from acidic A
deposition are reported in the literature, but some potential •
problems are identified. In New York State water from the Hinkley
reservoir has become acidified to such an extent, that lead
concentrations in drinking water at the tap exceed the maximum levels •
for human use (50 Pg/L) recommended by the New York State Department •
of Health (Turk and Peters 1978).
Fuhs and Olsen (1979) investigated drinking water in the Adirondack •
region of New York State. They found high metal concentrationns at
several test sites. At one home with a water pH of 5.71, copper and ^
lead levels reached 6.6 and 0.10 mg/L respectively in a water line 1
which had not been used overnight. In another home which had a water ^
pH of 4.95, copper concentrations of 2.3 mg/L were recorded from a
flushed line. Both of these homes obtained their water from shallow •
wells. The corrosive nature of the groundwater was estimated using |
the Langelier Saturation Index. The results indicate that a large
portion of the water in the area is corrosive. From their study, •
Fuhs and Olsen (1979) concluded that high concentrations of metals •
are present in homes with metal piping especially if the lines are
used intermittently.
Taylor (1982) has recently reported some of the data obtained from ™
an examination of surface and groundwater drinking water supplies in
New England. Utilizing the calcium saturation index and the tt
aggressive index as indicators of water quality, he concluded that |
raw water supplies in the region were generally very corrosive. In
addition, many of the watersheds tested had a very limited capacity im
to withstand further acid input without deteriorating further. An •
analysis of past water quality records for the area is proposed to
determine what increment of the acidic conditions may be attributed
to acidic deposition. fl
A recent study completed for the Department of National Health and
Welfare (Meranger and Khan 1982), measured the leaching rates of
metals from the plumbing systems of cottages in central Ontario on
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5-7
acid-sensitive lakes. In addition, samples of tap water taken from
flushed and unflushed systems were analyzed for Cu, Pb, Zn and Cd.
Results from the leaching study indicate that the maximum rate of
leaching occurs during the first 2 hours of water residence in the
system, although levels do rise for up to 10 days. The maximum
concentration of metals observed in the study resulted from 10-day
static samples. Maximum values of 4.56 mg/L, 0.478 mg/L, 3.61 mg/L
and 0.0012 mg/L were recorded for Cu, Pb, Zn and Cd respectively.
The concentrations of all metals decreased after flushing but still
remained above source levels. This indicates that the concentration
of metals are related to the contact time in the distribution
system.
The cottage survey consisted of obtaining tap water samples from
standing and flushed supply systems for metal analysis. The median
values recorded for static samples were 0.067 mg/L Cu, 0.014 mg/L Pb,
0.219 mg/L Zn and 0.0002 mg/L Cd. As anticipated, the metal
concentrations in the water decreased by up to 80% following
flushing. There was only one recorded instance where a sample
slightly exceeded federal guidelines (0.053 mg/L Pb), and this value
was obtained from a standing water supply.
Based on these preliminary data, no immediate threat to human health
is perceived. Careful monitoring of the situation should continue to
document any significant alterations in metal levels that may occur
in the future.
Consumption of drinking water with a low pH from municipal sources is
not a major issue. The raw water utilized by the treatment plant is
adjusted to drinking water standards by the addition of appropriate
substances. The only health concern related to this matter is to
ensure that no excessive accumulation of cations (e.g., Ca2+)
develops as a result of the neutralization process.
Groundwater may also become acidified in poorly buffered areas
(Cronan and Schofield 1979). For example, in Sweden some well water
became so acidified that substantial corrosion of household plumbing
occurred (Hultberg and Wenblad 1980). An occurrence of this kind
could lead to increased levels of such metals as aluminum, zinc,
copper, lead and cadmium in drinking water.
Many wells in the Precambrian area of Ontario are located in
proximity to shallow bedrock, so the potential for acidification
exists. The first field surveys were carried out in 1980 in the
Muskoka-Haliburton area. A total of 85 groundwater samples were
field-tested, and 28 samples were analyzed in the laboratory for
major ions and some trace metals. Groundwater was sampled in July
from shallow springs and wells from both bedrock and overburden
formations. Eleven of the 85 samples had pH values less than 6.0
with the lowest value being 5.2. October sampling of five of the low
pH wells resulted in only one sample with a pH value less than 6.0,
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5-8
5.1.4 Recreational Activities in Acidified Water
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suggesting that groundwater pH may fluctuate during the year. The
lowest recorded pH of 5.2 was from a shallow well servicing a
permanent home near Bracebridge, Ontario. 4
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5.1.3 Drinking Water From Cisterns
Sharpe et al. (1980) provide significant information on the effects
of deposition of lead and cadmium as well as acidic deposition on M
the quality of drinking water from cisterns in Clarion County, V
Pennsylvania. Wet and dry deposition of lead and cadmium resulted in ™
solutions which were in the same order of magnitude as recommended
United States drinking water limits (50 Ug/L and 10 Ug/L I
respectively). Lead levels in tap water from cisterns were much ™
higher than those found in the source water; about 16% of the
households sampled had tap water levels in excess of the United •
States drinking water standard. The investigators concluded that the f
increase in tap water lead levels resulted from acid corrosion of the
lead soldered joints in the cistern and plumbing. Thus, cistern _
water users are at special risk in areas of high acidic deposition. •
There is a time dependence for the initiation of adverse health
effects resulting from drinking water contaminated with metals at, or •
approaching, the concentrations listed in Table 5-1. For example, V
brief episodic excursions of lead over the recommended standard
associated with snowmelt derived acidity in water from small lakes or *|
streams, is not likely to be of major concern. Longer or continual •
consumption of water containing lead levels 25 pg/L could be of
concern (NAS 1977c), although the actual standard for drinking water
lead levels in the United States and Canada is 50yg/L. •
I
Due to the increased atmospheric fallout of acids, poorly buffered
bodies of water have shown a decline in pH. Some of these lakes and ,_
rivers have attained acidity levels in excess of the federal •
guidelines for drinking water of pH 6.5 (NHW 1980). As a result of *
this situation, attention has been focused on the safety of these
affected waters. There is concern that recreational activities in •
these waters (e.g. , swimming) may prove to be detrimental to human •
health.
However, the rationale for the federal pH standard, which was ^
accepted by most provinces in Canada, is based on corrosion and
incrustation effects, not on health considerations. Generally metal
corrosion may become a problem when the water pH falls below 6.5 and ,1
scale build-up on supply systems is usually encountered above pH 8.5. «
If an organ system was susceptible to the effects of acidified water,
then it is felt that the eye would be the most likely candidate. •
This possibility is rather remote at best. |
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5-9
TABLE 5-1. CANADIAN AND UNITED STATES DRINKING WATER
GUIDELINES FOR TOXIC METALS ( yg/L)
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Canadian United States
_ Lead 50 50
V Mercury 1 2
™ Cadmium 5 10
Copper 1000 1000
• Zinc 5000 5000
• Arsenic 50 50
Selenium 10 10
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5-10
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Doull et al. (1980) states that acid-induced eye damage is a function
of pH and the capacity of the anion in question to combine with «
protein. Therefore the clinical findings may vary depending on the •
concentration and type of acid under consideration. Exposures to '
mixtures of acids would be even more complicated to assess. With
hydrochloric acid virtually no clinically significant effects are M
present above pH 3 while some discomfort was evident between pH 4.5 •
and 3.5. Doull et al. (1980) and Grant (1974) reported that in
normal rabbit eyes, only acid solutions below pH 2.5 produced any M
significant eye injury and in humans brief contact of solutions from J
pH 7 to as low as pH 2 caused no damage. However, the subjects did
complain of "an increasingly strong stinging sensation" as the acid ^
level increased. •
Due to the lack of scientific data on this issue, the Department of
National Health and Welfare (Canada) is funding research into •
investigating the effects of acidified water on the eye. The water W
used in this work will be obtained from some of the most severely
affected lakes in central Ontario to simulate the worst possible 4
conditions a person may be subjected to. Results from this study •
are anticipated in mid-1982. A preliminary study by Basu (1981) has
indicated that there are no ocular clinical effects produced by ^
short-term exposures to lake water with pH values as low as 4.6. I
Based on all of the above information it would seem unlikely that any ™
ocular exposure to the mildly acidic waters in affected regions would
produce any harmful health effects. However, a final conclusive •
statement on this matter should be reserved until the results of the |
Health and Welfare study are complete.
5.1.5 Direct Effects; Inhalation of Key Substances Related
to Long Range Transport of Air Pollutants
Although no direct health effects were associated with acidic B
deposition per se, deleterious effects have long been attributed to
inhalation of high concentrations of several important pollutant ^
classes that were implicated as precursors to acidic deposition. £
These include ozone and other photochemical oxidants, acidic aerosols
and other particulate matter, and oxides of sulphur and nitrogen. A
The effects associated with these pollutants led to the establish- •
ment of minimum air quality standards for each pollutant in both the
U.S. and Canada (Table 5-2). Extensive reviews of the health effects
literature were recently conducted for these pollutants (e.g., NAS jl
1977a,b, 1978; USEPA 1978, 1980, 1981, 1982; WHO 1979). These •
reviews generally support the notion that attainment of the respec-
tive air quality standards will protect public health. The reader is ||
referred to these documents for a comprehensive assessment of the J|
effects literature. The discussion below is intended only as a brief
summary of some aspects of interest. —
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5-12
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Ozone is a secondary gaseous pollutant, formed as a result of
photochemical reactions of volatile organic chemicals and nitrogen
oxides. Therefore, the formation and transportation of ozone is V
limited by the production of NOX and suitable environmental Q
conditions.
Ozone is a deep lung irritant, capable of causing death from •
pulmonary edema in serious cases. Sublethal exposures produce
substernal tightness, irritation of mucous membranes, dry cough, and
headache. Morphological and biochemical changes in the lung are also 9
observed following exposures to low levels of 03. In addition, ™
extrapulmonary effects were documented. Changes in the circulating
red cells of animals exposed to ozone were reported. Since 03 is I
unlikely to penetrate the pulmonary mucosa, the alterations are f
suggested to arise as a result of a chain reaction when various free
radicals are formed (Goldstein 1980). •
The effects of ozone are influenced by factors other than
concentrations and length of exposure and young animals are more
susceptible. Elevated temperatures, increased relative humidity and f
exercise all increase the toxicity of 03 (Doull et al. 1980). As 9
with other toxicants, the individual health of a person has an effect
on the results. People with respiratory disease (e.g., asthma, JM
emphysema or bronchitis) are believed to be particulary sensitive to £
low-level exposures.
Ozone produces eye, nose and throat irritation in the 0.1 - 0.15 ppm •
range (Ferris 1978), and the America Lung Association (1977) states *
that significant health effects are found when ozone levels are above
0.37 ppm. Goldsmith and Nadel (1969) found consistent increases in •
airway resistance after 1-hour exposures of 1.0 ppm, while other •
researchers found similar results with lower 03 concentrations and
longer exposure times (Kerr et al. 1975; Young et al. 1964). •
There is some controversy surrounding the possible synergistic
properties of 03 in combination with other pollutants. Although ^
evidence was developed that suggested ozone has an enhanced toxicity M
when combined with S0£ (Hazucha and Bates 1975; NRC 1975), recent ™
studies have found no clear support for such synergism (Bedi et al.
1979; Kleinman et al. 1981). fi|
One theory gaining some acceptance deals with exposures to low
concentrations of ozone over several days. Ozone is thought to have •
a short-term cumulative effect above a threshold value (Folinsbee •
et al. 1980). In consecutive exposure studies a decrease in
pulmonary function is maximized during day 2-3. Additional exposures ^
result in a return to pre-exposure values or an improvement in •
pulmonary response on day 4-5 (Farrell et al. 1979; Hackney et al. ™
1977). These results also support the notion that some form of
adaptation or tolerance is developed following repeated exposures. •
This tolerance, however, may be of limited duration and dependent on |
peak concentration. All of these experiments have dealt only with
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5-13
short-term responses, and the consequences of biological changes
resulting in tolerance are not known. Specifically, more data is
required on chronic or long-term exposures in view of,animal studies
suggesting increased susceptibility to infection, morphological
abnormalities and extrapulmonary effects (USEPA 1978).
In summary, brief exposures to ozone can be linked to alterations in
pulmonary function. At low levels, 03 produces eye, nose and
throat irritation. Based on the available quantitative evidence, the
established air quality standards appear to be protective of public
health. Ozone levels achieve maximum readings from April to July and
the annual mean concentration exceeds maximum acceptable levels at
several locations, predominantly in eastern Canada (see Tables 4-6,
4-7). The ozone air quality standards are also exceeded in a number
of northern U.S. cities (Table 4-5).
Nitrogen oxides, among the photochemical irritants, are primarily
derived from internal combustion engines, and NO is rapidly converted
to M>2 in the atmosphere. A recent national air pollution survey
(Environment Canada 1979) indicates that N0£ concentrations are
highest from January to July, but concentrations do not exceed
existing Canadian guidelines. Nitrogen dioxide, like ozone, is a
deep lung irritant and can produce pulmonary edema in severe cases.
Both short- and long-term exposures to N0£ enhance susceptibility
to infections. There is some evidence that elevated levels of N0£
will produce an increase in respiratory disease (Florey et al. 1979;
Guidotti 1978; Speizer et al. 1980), but presently it is not clear
whether transient peaks of N0£ or long-term exposure to low levels
are primarily responsible for these observations. The levels of
N02 in Canada are relatively low and unless further information
becomes available, current standards seem adequate to protect human
health.
Sulphur oxides (S02) and related particulates are also respiratory
irritants. An important feature of sulphur oxides are their ability
to interact with other pollutants to form substances that vary in
their respiratory instancy potential. The most prominent response to
inhaled S0£ is bronchial constriction leading to increases in flow
resistance. Healthy individuals begin to respond to S02 peaks of
about 5 ppm while sensitive individuals may respond to short-term
exposures of less than 1 ppm (USEPA 1981).
Various sulphate forms have also been associated with increases in
pulmonary disease and some concern has been raised over statements
linking airborne sulphates to human morbidity/mortality (Lave and
Seskin 1973). This and other investigations were reviewed by the
subjects and from ill-defined groups that may have confounding
factors (e.g., smoking habits and health problems) and may lack good
exposure estimates and contain uncertainties with respect to
statistical models. This limits the use of these studies for
assessing health effects of specific pollutants. At best, they
provide qualitative support for the association of sulphur-
particulate pollution with health effects (USEPA 1981).
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5-14
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As with nitrogen oxides, the levels of SC>2 asssociated with
transboundary transport between the U.S. and Canada are relatively
low, and current standards seem adequate to protect human health. I
However, some concern has been expressed about regional transport of •
sulphur containing fine particles including sulphuric acid, sulphates
and associated substances. The effects of general particulate matter •
and sulphur oxides were reviewed recently by Ware et al. (1981) and •
Holland et al. (1979) and are the subject of a revised EPA criteria
document (USEPA 1981) and staff paper (USEPA 1982). These reviews _
suggest that it would be inappropriate to single out sulphates as the •
only significant component of the sulphur - particle complex. Until •
ongoing standard reviews and perhaps additional research is con-
ducted, it appears that attainment and maintenance of the current I
U.S. and Canadian standards for particulate matter would provide |
reasonable public health protection. The U.S. is considering the
possibility of new standards based on particle size. No changes in A
the maximum acceptable levels for suspended particulate matter are •
anticipated in Canada. *
In summary, to the extent that transboundary transport contributes •
significantly to violations of the air quality standards listed in 0
Table 5-2 (an issue for Work Group III to resolve), the matter should
be the subject of the bilateral discussions. fij
5.1.6 Sensitive Areas and Populations at Risk - Health ,_
Certain areas are sensitive to acidic deposition resulting in ™
contamination of fish and drinking water supplies. These include
areas with poorly buffered lakes and streams (with a viable fish 11
population), watersheds with unusual accumulations of metals in •
sediments or soils, areas which lack drinking water treatment
facilities, and areas with substantial lead plumbing. m
Some populations are more susceptible to environmental insults than
others. These populations include those dependent on fish from
acidified waters as a major dietary staple, those with elevated •
mercury or lead blood levels from other exposures, those dependent on W
cisterns as a primary source of drinking water, and women of
childbearing age as well as children.
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5.1.7 Research Needs
Due to the common areas of interest between the health effects and
aquatic sections, there are also similar gaps in data bases and
research requirements. For example, work is required in the •
following areas: •
1. Acidic deposition appears to increase the mobilization of •
metals from soils and the leaching rates in water distribution •
systems. Therefore, data is needed to further clarify the
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5-15
levels, species, and temporal variations of these metals and
their effects on biological systems.
2. A data base needs to be developed that will enable researchers
to identify sensitive areas and receptors in order to predict
which regions have a greater risk of developing health related
problems.
3. There is an apparent relationship between declining water pH
and increasing Hg levels in the food chain. Therefore, more
data related to their exposure levels and Hg content in various
species is needed.
4. More research is needed to differentiate the health effects
resulting from short-term exposures and long-term exposures to
pollutants subjected to long range transport.
5. Carefully designed and controlled epidemiological studies are
necessary in order to relate exposure to various air pollutants
to the health of susceptible individuals as well as the general
population.
6. The relative contributions of transported air pollutants that
contribute to acidic deposition should be determined. The
levels of these pollutants should be compared to the National
Ambient Air Quality Standards.
5.2 VISIBILITY
This discussion is largely adapted from a recent EPA staff assessment
(USEPA 1982). The effect of transboundary pollution on visibility is
directly related to air quality, rather than deposition. The
particulate phase precursors to acidic deposition (mostly sulphuric
acid aerosol and various ammonium sulphate aerosols) as well as other
fine particles play a major role in atmospheric visibility.
Available data suggest that nitrates exist predominantly in the
vapour phase and are for the most part of little consequence to
visibility in eastern North America. For some isolated point
sources, however, NC>2 may produce visible brown plumes at distances
of 100 km from the source (Menlo 1980).
5.2.1 Categories and Extent of Perceived Effects
Impairment of visibility is perhaps the most noticeable and best
documented effect of particles in current North American atmospheres.
It is often equated with "visual range" as measured by airport
weather observers. However, visibility in a broader context relates
to visual perception of the environment and involves colour and
contrast of viewed objects and sky, atmospheric clarity, and the
psychophysics of the eye-brain system (USEPA 1979).
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5-16
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I
For present purposes, it is useful to classify pollution-derived
effects on visibility into two categories: (1) regional haze, and
(2) visible plumes. The nature and extent of these effects are I
determined largely by the distribution and characteristics of •
anthropogenic and natural particles and, to a lesser extent, by
N02• Salient features of both categories are outlined below. •
Regional haze is relatively homogeneous, reduces visibility in every
direction from the observer, and can occur on a geographic scale _
ranging from an urban area to multistate regions. Increased haze •
reduces contrast causing more distant objects to disappear. Nearby ™
objects can appear "flattened" and discoloured, the horizon sky is
whitened, and scattered light is perceived as a gray or brown haze B
(Charlson et al. 1978). When urban light and haze combine at night, |
the contrast between the night sky and the stars is reduced, markedly
limiting the number of stars visible in the night sky (Leonard et al. •
1977). |
The best available indication of the extent and intensity of regional
haze with time is visibility (visual range) data routinely measured •
at airports and some other locations. Some uncertainties arise from "
the use of such data to characterize regional visibility; among them
are differences in target quality and observers between sites and at jfl
the same site, representatives of the airport location, and potential j|
biases in measurement techniques. Analyses of airport visibility
trends from 1948 to 1974 suggest that visibility in the eastern U.S. _
declined over that period, particularly during the summer months •
(Husar et al. 1980; Trijonis et al. 1978b). The analysis of *
visibility trends has recently been extended for this report by Husar
and co-workers (Husar pers. comm.) to include Canadian and U.S. data •
through 1980. Figure 5-3 presents the results of the preliminary •
analysis. This figure represents extinction weighted airport
visibilities for about 300 U.S. sites and 177 Canadian sites (94 in •
eastern Canada). Sites were selected on the basis of a reasonably •
continuous record. The airport data reflect 5-year quarterly means
of noon extinction, (3.9/visibility) from 1950 to 1980 exclusive of
readings with fog, precipitation or blowing material. Because the •
figure represents average extinction, the estimated visibility as W
derived from the indicated scale is weighted to lower than actual
average visibility on fog/precipitation free days. These data are -M
quite preliminary and subject to the usual caveats regarding airport l|
visibility trends. The U.S. results appear consistent with other
published data. The lower density and variability of Canadian sites ^
make regional representations presented considerably less reliable. •
Differences exist between some features in the figure and other *
examinations of specific sites in Canada; these may be related to
site or to different treatment. Until further examination of the I
Canadian data is completed, the results should be treated with •
caution (Christe pers. comm.).
The figures show eastern visibility is substantially less than that •
in the west. The unusual area of persistent low visibility in
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QTR 2
QTR 3
1/km
0.18-0.24
0.24-0.30
0.30-0.36
> 0.36
QTR 4
ENA 0.24
EUS 0.26
SURE 0.27
Figure 5-3a. Seasonal and spatial distribution of long-term trends
in extinction - weighted airport visibilities for North
America, 1950-54 (after Husar pers. comm.).
-------
QTR 3
0.24-0.30
0.30-0.36
> 0.36
Figure 5-3b. Seasonal and spatial distribution of long-term trends
in extinction - weighted airport visibilities for North
America, 1960-64 (after Husar pers. cornrn.)-
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\ o.
QTR 3
ENA 0.28
BUS 0.32
SURE 0.33
QTR 2
QTR 4
Figure 5-3c. Seasonal and spatial distribution of long-term trends
in extinction - weighted airport visibilities for North
America, 1970-74 (after Husar pers. comm.).
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5-20
QTR 3
ENA 0.26
EUS 0.30
SURE 0.31
Figure 5-3d. Seasonal and spatial distribution of long-term trends
in extinction - weighted airport visibilities for North
America, 1976-80 (after Husar pers. comm.)»
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5-21
northern most Canada has not been explained, but may be related to
very high humidity and the presence of ice crystals and, in
wintertime, reduced daylight. It is unlikely to be related in any
substantial way to anthropogenic sources. Moreover, the number of
sites in this large region is quite limited, and may not be truely
representative of precipitation/fog free days.
Consistent with earlier U.S. work, Figures 5-3a,b and c shows that
while visibility in some urban areas improved or stayed the same from
1950-74, the occurrence of episodic regional haze appears to have
increased in the eastern U.S. and portions of southern and eastern
Canada. Summertime (Quarter 3) trends are most pronounced in these
areas. Since 1972, regional visibility in eastern Canada (Figure
5-3d) and in both the east and west of the U.S. apparently has
improved slightly but not to pre-1960 levels (Marians and Trijonis
1979; Sloane 1980). Whether this recent improvement is related to
more favourable meteorology or reduced regional particle and sulphur
oxide emissions is not known with certainty, but such reductions are
reflected in emissions inventories in both east and west.
Regional differences in average U.S. visibility are illustrated in
Figures 5-4 a and b. As indicated by suburban and nonurban airport
data, visibilities in the east are substantially lower than in most
of the west. Some of the differences between east and west may be
related to the lower regional humidity in the west, but a more
important difference is the generally higher regional particle
loading in the east. Based on: (1) long-term historical data in the
northeast from 1889 to 1950 (Husar and Holloway 1981);
(2) examination of airport visibility trends after deleting data
possibly influenced by obvious natural sources (i.e., fog,
precipitation and blowing dust) (Figure 5-3); and (3) current
assessments of natural sulphur sources and regional fine particles
levels (Ferman et at. 1981; Galloway and Whelpdale 1980; Pierson
et al. 1980; Stevens et al. 1980), anthropogenic particulate
pollution would appear to dominate eastern regional haze. Relying on
the analysis of Ferman et al. (1981), it has been estimated that in
the absence of anthropogenic sources summertime visibility in the
Shenandoah Valley would range between 60 and 80 km (36-50 miles)
(USEPA 1981). The median daytime visual range actually observed
during the 1-month study at this site was four to five times lower (9
miles).
Visible plumes of smoke, dust, or coloured gas obscure the sky or
horizon relatively near their source of emission (USEPA 1979).
Black, gray, or bluish plumes are caused by particles. Brownish
plumes may be caused by N02 or particles. Perception of plumes
(and regional haze) is strongly influenced by factors such as viewing
angle, sun angle, and background objects (USEPA 1981). Because
visible particle plumes often are subject to opacity regulations and
because they are usually quite localized in character, the focus here
will be on urban and larger scale regional haze.
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Figure 5-4.
10"
Median 1974-76 visibilities (miles) and visibility
isopleths for suburban/nonurban airports: (a) yearly,
and (b) summertime (Trijonis and Shapland 1979). Data
are subject to uncertainties associated with suburban
airport observations, but show general regional
patterns. The clear differences between east and west
are parallelled by regional humidity and nonurban fine
particle levels. Summertime fine mass averaged from 22
to 25 yg/m3 at 12 eastern nonurban sites (Watson
et al. 1981) and about 4 pg/m3 for 40 Rocky Mountain
and southwest background sites (Snelling 1981).
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5-23
5.2.2 Evaluation of Visibility
Visibility impairment may adversely affect public welfare in
essentially two areas: (1) the subjective enjoyment of the
environment (aesthetics, personal comfort and well-being), and
(2) transportation operations. The aesthetic aspects of visibility
values can be categorized according to: (1) social/political
criteria, community opinions and attitudes held in common about
visibility; (2) economic criteria, the dollar cost/benefit associated
with visibility; and (3) psychological criteria, the individual needs
and benefits resulting from visibility. These categories are not
exclusive, but relate to different approaches for measuring somewhat
intangible values. Evidence on visibility effects is drawn from
studies of social perception and awareness of air pollution, economic
studies, and visibility/air transportation requirements. These
studies are summarized in Table 5-3 and evaluated briefly below.
5.2.2.1 Aesthetic Effects
Assessment of the social, economic and psychological value of various
levels of visibility is difficult. The criteria document, an EPA
report to Congress (USEPA 1979), Rowe and Chestnut (1981), and Fox
et al. (1979), discusses and evaluates several approaches that have
been used or proposed towards this end. In particular, preliminary
studies of social awareness/perception and the economic value of
visibility in urban and nonurban areas support the notion that
visibility is an important value in both settings.
Early social awareness studies (DHEW 1969; Schusky 1966; Wall 1973)
conducted in polluted urban areas have generally shown that at higher
pollution levels an increasing portion of the population is aware of
air pollution and considers it a nuisance. In St. Louis, a linear
relationship was observed between annual particle levels (50-200
yg/m TSP) and public awareness and concern. At 80 ug/m annual
geometric mean TSP, about 10% of those surveyed reported air
pollution as a nuisance (Schusky 1966). Although it is reasonable to
attribute more of these and other perception results to particulate
matter than to gaseous pollutants (Barker 1976; Wall 1973), the
relative importance of visibility degradation by plumes and haze as
compared to dustfall was not clearly addressed in these studies. A
more recent study of perception of air pollution in Los Angeles
(Flachsbart and Phillips 1980) represents the most comprehensive
evaluation of major pollutant indicators and perception to date.
Five gaseous pollutants (0^, CO, N02, NO, S02), four particle
related indices (TSP, dustfall, CoH, visibility) and six averaging
times were compared with perceived air quality as reported by 475
respondents living in 22 residential areas in Los Angeles County.
Only two indices, visibility and ozone, were consistently
significantly related ( a = 0.001) to perceived air quality for all
averaging times. The highest correlation coefficient (Kendall's T )
occurred for yearly visibility ( T= -0.29) and for number of days
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5-24
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5-25
visibility was less than 3 miles ( T = 0.32).a Of the other
particle indices, only monthly average dustfall was significantly
correlated ( T = 0.12) with perception. Consistent with other
studies, the survey also found that air quality is valued less by
minority groups and long-time urban residents than by whites (all
income classes) and those with some history of rural residence.
The two major approaches to economic valuation of visibility include:
(1) survey (e.g., iterative bidding), and (2) property value studies.
The major published iterative bidding studies of visibility,
conducted in the rural southwest and in the Los Angeles area (South
Coast Air Basin), are summarized and evaluated in Table 5-4. The
Four Corners and Lake Powell studies deal only with single sources
and visible plumes, while the Farmington and Los Angeles studies
address haze. The preliminary nature of these studies makes them
useful primarily as qualitative indicators of the economic value of
visibility. Among the more important limitations of the published
results are the following:
1. None of these studies has measured existence values (benefit of
just knowing pristine areas exist, regardless of intent to use
them) or options values (wish to preserve the opportunity to
view an unimpaired vista). Rowe and Chestnut (1981) suggest
that existence values of good visibility in natural settings may
be significantly greater than measured activity values.
2. The studies may be subject in varying degrees to methodological
problems. The Farmington study, in particular, discovered a
number of biases probably related to the credibility of the
contingent market. These biases were not always large but show
the difficulty of valuing visibility through iterative bidding.
3. Even if the available results could be taken at face value, so
few studies have been conducted that results cannot be directly
transferred to other areas of the country. For example, it
might be expected that willingness to pay for improved
visibility in Los Angeles might be greater than that for areas
in flat terrain without background hills or mountains.
Despite their limitations, the iterative bidding studies suggest that
visibility is of substantial economic value in both urban and natural
settings. Although the value of visibility in other areas may vary
significantly from that suggested by studies in the rural southwest
and Los Angeles, no a priori reason exists to suggest that visibility
is of little value in heavily populated eastern urban and rural
areas. With respect to recreational settings, of the 23 most heavily
used national parks and monuments, 11 were in the East (NFS 1981).
In 1979 over 90 million visits were recorded in all eastern National
Park Service managed facilities.
a Kendall's T is a nonparametric statistic. The negative
correlation between the number of people perceiving poorer air
quality and visibility and the positive correlation with the number
of days visibility is less than 3 miles are consistent with
expectations.
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5-26
TABLE 5-4. SUMMARY RESULTS OF ITERATIVE BIDDING VISIBILITY STUDIES (after Rowe and Chestnut 1982)
I
Design Elements/Study (Year)
Total Interviews Conducted
(Percents Usable)
Scenarios
Four Corners
NM (1972)b
1,099 (69?)
Emissions, strip
minings, trans-
Lake Powel I
UT/AZ (1975)c
104 (79?)
Air Qua I ity and
Power Plant
Farmington
NM (1977)d
130 (92%)
Air Qua 1 ity and
Power Plant
South Coast
Air Bas in
CA (1978)e
345 (NR)
Air Quality
and Health
-•
1
mission lines,
single power plant
Payment Vehicle
Wi 1 1 ingness to Pay Bid
Comparisons
a) Yearly bids for
individual resi-
dents households
b) Dai ly bid for
individual rec-
reation ists min.
$1.00
c) Aggregate Yearly
Value
Tested for bias/found
biases
Comment on Total Values
Capital letters refer to
for going from scenario A
b Randal 1 et al . 1974
<: Brookshire et at. 1976
d Rowe et al. 1980
e Brookshire et al. 1979
A-worst
B-midd le
C-best
User fees/
Electricity Bill
A-Ca $85
B-C $50
A-B
~
A-B $15.54 M
A-C $24.57 M
No/No
Values for total
affected four
corners region;
attributed mostly
to part iculate air
pol 1 ution reduc-
tion at single
source, but
difficult to
separate from
other visible
factors due to un-
standardized
scenarios.
Measured activity
values only.
scenarios listed above
(worst) to scenario C
A-No plant
B-Plant, no plume
C-Plant with plume
User Fees
__
A-C $2.95
A-B $.414 M
A-C $.727 M
No /No
Examined one of
fifteen potentially
affected parks;
pictoral represent-
ations not con-
sistent across A-C.
Measured activity
val ues onl y.
. Thus "A-C" in the
(best).
A-Visibi lity 120 km
B-Visibi lity 80 km
C-Visibi lity 44 km
Utility Bil I/
Payrol 1 Reduction
User Fee
A-Da $82
A-B $57
B-D $43
A-D $2 .44
A-B $1.47 M
A-C $1 .99 M
A-D $2.14 M
B-C $1.1 M
YesAes
Val ue for San Juan
County, New Mexico
and Mavajo recrea-
tion area only.
Affected area up to
10 times larger
strategic bias not
found, but other
bias problems with
contingent market
found. Measured
activity values
only.
A-Poor Visibi lity-W
2 mi les
B-Fair Visibi lity-^
12 miles "•
C-Good Visibi lity«
28 miles
Utility Bil I/ •
Monthly Payments »
to Conservation
Fund ^.
1
•
Pi
A-Ca $245 •
B-C $243 •
A-B $174
1
w
30? i mprovement •
•
$.58 - $.65 B
res/ res
About one-th ird
of benefits are
related to •
aesthetics, 1
two-thirds to •
health. Resu Its
reasonably con- _
sistent with •
companion pro- £
perty value
study. Measured
user values. •
Some question- B
naire design 9
biases related
to aesthetics. —
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Four Corners case refers to the bids
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A large number of property value studies related to air quality have
been conducted. These have been reviewed by Freeman (1979a,b) and
Rowe and Chestnut (1981). Although a variety of air quality
indicators have been used, the results of awareness/perception
studies strongly suggest visibility plays an important role in air
quality related impacts on property values (Rowe and Chestnut 1981).
This contention is best documented in the case of the South Coast Air
Basin property value survey (Brookshire et al. 1979). There, the
estimated annual benefit of a 25-30% improvement in air quality based
on property values was about $500 U.S. per household. These results
are qualitatively similar to the companion iterative bidding study
(about $300 U.S./household). The bidding study suggests that 22-55%
of the bids to improve visibility were related to aesthetic effects.
Both the bidding results and the perception study (Flachsbart and
Phillips 1980) conducted in the same area support the possibility of
substantial impacts of visibility on Los Angeles area property
values.
None of the other published property value studies are accompanied by
companion studies that suggest what portion willingness to pay for
improved air quality may be due to visibility. Moreover, theoretical
problems remain in relating willingness to pay functions from
property value differentials (Rowe and Chestnut 1981). No single
study has examined all of the variables that might be important in
influencing property values. Because air pollution tends to be a
small influence compared to other variables, earlier studies that
examined a limited number of variables are particularly suspect. In
essence, the available literature suggests that perceived air
pollution, and hence visibility, may have tangible effects on
property values in urban areas such as Washington, Boston, Los
Angeles, and Denver (Rowe and Chestnut 1981); nevertheless,
additional theoretical and empirical work is needed before reliable
and transferable quantitative relationships for visibility evaluation
can be established.
5.2.2.2 Transportation Effects
Although all forms of transport may be affected by poor visibility
(e.g., slowing of highway traffic by anthropogenically induced fog),
at current ambient levels, aircraft operations appear to be most
sensitive. When visibility drops below 3 miles, both U.S. FAA and
Canadian safety regulations restrict flight in controlled air spaces
to those aircraft and pilots that are certified for operation under
Instrument Flight Rules (IFR) (FAA 1980a). The most severe impact in
such cases is usually on non-IFR general aviation aircraft which are,
in effect, grounded or forced to search for alternate landing sites.
In 1979, there were about 210,000 active general aviation aircraft
which accounted for about 84% of total airport operations in the U.S.
Over 23% of all general aviation aircraft had no IFR capability
(Schwenk 1981). Estimates of the percentage of pilots certified for
IFR in the U.S. and Canada are not available. Commercial, military,
and other IFR aircraft operation also may be affected by reduced
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visibility. Under IFR conditions, the number of arrivals and depar-
tures per hour can be significantly decreased as compared to Visual
Flight Rules (VFR) conditions. The effect varies with airport, and •
in some cases, the visual range cutoff for the most efficient visual |
approaches (VAPs) may be 5 miles (FAA 1980b). For example, the per-
formance standard for one configuration at Boston Logan International «
Airport is 109 operations per hour for VAPs (5 miles), 79 operations •
per hour for "basic" VFR (3-5 miles), 79 operations per hour for ™
"controllers" visual approach IFR (2-3 miles), and 60 operations per
hour for "category I" IFR (2 miles). Thus, depending on airport I
configuration, schedules, and the extent and duration of haze induced m
visibility reduction, delays in commercial and other aircraft opera-
tions can occur. In large segments of the eastern U.S., midday •
visibilities less than 3 miles with no obvious natural causes occur •
2-12% of the days in the summer and 1-5% of the time during other
seasons (USEPA 1981). Visibilities less than 5 miles would, of _
course, be more frequent. Based on typical eastern summertime H
diurnal cycles in humidity and light scattering (e.g., Ferman et al. ™
1981), the occurrence of morning visibilities (6-8 a.m.) less than
3-5 miles would be somewhat greater than for midday visibility, even A
discounting naturally occurring fog. |
Compared with other modes of transport, air travel is generally «
considered to be safe. It is not, however, riskfree; based on •
reasonable expectations and the available record, air pollution
visibility impairment would tend to increase risks of aircraft
operation (U.S. Senate 1963). Failure to see and avoid objects and •
obstructions during flight is one of the ten most frequent cause W
factors for general aviation accidents (FAA 1978). Another important
cause factor is continued VFR flying into adverse weather. Although •
such action is normally attributed to errors in judgment (FAA 1978), |
in some cases, pilots who by choice or necessity fly in the mixing
layer, could fail to distinguish storm fronts or thunderclouds from ^
the prevailing haze until they are upon the adverse weather. The •
available data in the criteria document do not, however, permit any ^
quantitative assessment of the risks to commercial and general
aviation aircraft operation associated with reduced visibility. ft
The available information of the effects of visibility on
transportation suggests that episodic eastern regional haze tends to •
curtail substantial segments of general aviation aircraft and slow •
commercial, military, and other IFR operations on the order of 2-12%
of the time during the summer. The extent of any delays varies with
airport. Reduced visibility may also tend to increase risks •
associated with aircraft operations in the mixing layer, but ™
quantitative assessments are not available.
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5.2.3 Mechanisms and Quantitative Relationships
The mechanisms by which atmospheric pollutants degrade perceived •
visibility are reasonably well understood (Friedlander 1977;
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Middleton 1952). Visibility impairment is the result of light
scattering and absorption by the atmospheric aerosol (particles and
gases). The "extinction" or attenuation coefficient (oext) is
a measure of aerosol optical properties and is the sum of blue sky of
Rayleigh scattering by air molecules (tfRg), absorption by
pollutant gases (oag), and particle scattering (aSp) and
absorption ( aap). Visibility is inversely related to total
extinction from these sources. Blue sky scattering is relatively
constant and is significant only under relatively pristine
conditions. Absorption by pollutant gases, notably NC>2, usually
contributes only a small amount to total extinction (USEPA 1981).
Even brown hazes in Denver and Los Angeles formerly attributed solely
to N02, are often dominated by particles (Groblicki et al. 1980;
Husar and White 1976). Thus atmospheric extinction and visibility
impairment are normally controlled by particulate matter. Important
causes include natural sources (e.g., fog, dust, forest fires, sea
spray and biologic sources) and anthropogenic sources of sulphur
oxide, soot and other particles, nitrogen oxides, and volatile
organics (USEPA 1979).
Reduction of visual range by particle extinction is normally domin-
ated by fine particles.3 The only important exceptions are some
naturally occurring phenomena including precipitation, fog, and dust
storms, where larger particles control visibility. Theoretical
calculations show that extinction/unit mass efficiencies are
substantially greater for fine particles in the 0.1 to 2.0 size range
than for larger particles (Faxvog and Roessler 1978). For most
commonly observed size distributions of particulate matter, the
increased extinction efficiency of fine particles results in fine
particles accounting for most of total extinction even though they
are only a third or so of the total mass of particles (Latimer et al.
1978). This theoretical expectation is borne out by the unique
experiment of Patterson and Wagman (1977) where independent
measurement of light scattering and particle size distributions
verified the importance of fine articles in controlling scattering in
New York City. In addition, a number of experiments have found
consistently high correlation (0.8 to 0.98) between light scattering
and fine mass (USEPA 1981).
The relative importance of scattering and absorption as well as the
extinction efficiency per unit equilibrated mass (y ) of fine
particles varies with location. On large regional scales, about
80-95% of particle extinction is due to light scattering (Waggoner
et al. 1980; Wolff et al. 1980), with the remainder due to
absorption. In urban areas absorption may account for up to 50% of
particle extinction (Waggoner et al. 1980; Weiss et al. 1978). The
particle scattering efficiency/unit mass varies from about 3-5
at various sites, with higher values tending to occur in eastern
locations (USEPA 1981).
a For purposes of this document, fine particles include those
smaller than 2.5 ym AED.
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The variations in fine particle extinction noted above are due
largely to variations in size, chemical composition and to some
extent, relative humidity. Based on theoretical (Faxvog and Roessler •
1978) and empirical (e.g., Groblicki et al. 1980; Stevens et al. |
1980; Trijonis et al. 1978 a,b) results, two components, hygroscopic
sulphates and elemental carbon, generally tend to be most _
significant. Sulphate, with associated ammonium, and hydrogen ions •
and water, often dominates regional fine mass and extinction, ™
particularly in the East, while elemental carbon accounts for most of
the particle absorption observed in urban areas. The relative flj
importance of sulphates to extinction depends on relative humidity, I
both at the site in question and perhaps along the transport path
where secondary formation occurs. Project VISTTA found that •
sulphates formed in dry desert air were of relatively low light •
scattering efficiency, compared to sulphates apparently formed in
more humid conditions in southern California and transported to the _
desert (Macias et al. 1980). Our understanding of the role of fine •
organics and nitrates in light scattering is hindered by the lack of ™
reliable data. In the eastern regional haze these components are
likely to amount to less than half of the sulphate component, but may ft
dominate scattering in western urban areas such as Denver and ||
Portland (Cooper and Watson 1979; Groblicki et al. 1980). The
remainder of fine mass (soil-related elements, lead and trace •
species) contributes only a minor amount to extinction in most U.S. B
atmospheres (Stevens et al. 1978).
Humidity is important to visibility because of the presence of fine •
hygroscopic aerosols (e.g., sulphates) which tend to absorb ™
atmospheric water and thus increase light scattering. Measurements
in several areas suggest that the extinction due to fine particle •
scattering will increase by a factor of about two as relative |
humidity is increased from 70% to 90% (Covert et al. 1980). Based on
laboratory studies, reduction in humidity from 90% to 70% might not ^
produce corresponding decreases in scattering because of hysteresis fl
(Tang 1980). In essence, the hysteresis effect means that the *
aerosol may tend to retain water absorbed at higher humidities even
at lower relative humidities. This effect has not yet been •
demonstrated to occur in the ambient air. £
Through the Koschmieder equation, the extinction coefficient, •
measured or estimated from fine particle levels, may be used to •
estimate visual range (USEPA 1979). The relationship has the general
form:
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a ext
Where: V = the visual range, the distance at which a large
black object is just visible against the sky.
0ext = total extinction, the sum of light scattering
and absorption by air molecules, fine particles, and
N02.
K = a function of the intrinsic target brightness and
observer threshold contrast (E). E is a function of
the observer and of target size.
Although a number of factors may limit the applicability of this
relationship, for homogeneous pollution, reliable extinction
measurements, uniform illumination, large dark targets, and moderate
visual ranges, agreement between experiment and theory is rather
good. The correlation between visual range and the scattering
portion of extinction is typically on the order of 0.9 where
comparisons have been made (USEPA 1981).
This relationship depends, in part, on human perception of contrast
as well as target size and brightness. For a typical observer with a
reasonable time for observation and large black targets, a
"threshold" of 0.02 is commonly assumed with K = 3.9 (USEPA 1979,
1981). Empirical determinations of K have yielded somewhat lower
values, ranging from 1.7 to 3.6 for studies discussed in the EPA
criteria document (USEPA 1981). The most complete analysis (Ferman
et al. 1981) reported a value of 3.5 for well mixed periods. The
lower values likely arise from higher threshold contrasts, nonideal
targets, (too bright and/or too small), and the exclusion of
absorption estimates. The available data also suggest that reported
airport visibilities may significantly underestimate standard visual
range. Thus lower values of K may be more appropriate for matching
airport data with higher values for observations in natural setting.
The relationship between extinction due to dry particle scattering
and fine mass is sufficiently stable over a wide range of areas that
reasonably quantitative estimates of fine/particle visibility
relationships can be made where long-term relative humidity is low
(<70%) and particle absorption is small or otherwise known. For such
purposes, the Koschmieder relationship can be written as:
V = K (2)
CTag + aRg + FMC
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5-32
Where FMC = fine mass concentration, and
•Y = (°ap +°Sp)/FMC
a ag = absorption by gases (usually small in nonurban
areas)
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o ap = particle absorption coefficient (in units of
inverse distance; e.g., km~l) •
o" s p = particle scattering coefficient
° Rg = Rayleigh or "blue sky scattering" |
(Rg~ 0.12 km"1)
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Thus , with appropriate K and Y derived from available studies , visual
range can be estimated from fine mass. Although less certain, the
measurements of Covert et al. (1980) and regression relationships A
developed by a number of investigators can be used to estimate fine •
particle/visibility relationships for higher humidities and
hygroscopic aerosols. The criteria document indicates that to
correct for the humidity effect (as determined by heated •
nepholometers and equilibrated filters), the amount should be W
increased by a factor of about 1.5 at 80% RH, and about 2 at 90% RH.
The Koschmieder relationship strictly applies for short-term •>
observations. In estimating long-term (e.g., annual) average •
visibility from long-term fine mass data, the temporal distribution
of fine particle concentrations (e.g., lognormal) must be specified, ^
or median values used. •
Figure 5-5 presents fine particle/visual range relationships for
three cases selected as representative of the range of normal B
situations encountered in the eastern North American regional haze •
and in western urban areas:
1. Y= 3 m2/g; representative of a dry aerosol, (USEPA 1981) at •
. Absorption may be 10% of extinction where
mass =2.7 m^/g. This is close to typical
_
measurements in western areas but below most eastern data (USEPA •
1981). •
2. Y= 6 m^/g; representative of the same aerosol as in 1) at 90% fl
humidity, a sp increased by a factor of 2. w
3. Y= 10 m^/g; representative of the similar aerosol, but with •
absorption accounting for 40% of extinction. Such high •
absorption (predominantly associated with carbon) is likely only
in urban areas.
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50
40
30
z
<
DC
—
>
20
0
VISUAL RANGE=3.9/^ext
CASE 1 = AEROSOL <50%RH(3m2/g;
CASE 2= AEROSOL 90%RH (6m2/g)
CASE 3 = AEROSOL 90%RH(10m2/g
30
25
20
15 E
10
25 50 75 100
FINE MASS CONCENTRATION (MQ/ m3)
125
150
Figure 5-5.
Visual range as a function of fine mass concentration
(determined from equilibrated filter) and Y , assuming
the "standard" K = 3.9. Because K is commonly lower in
nonideal application, results from this relationship
should not be compared directly to airport visibility
data.
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5-34
1. Analyses of the contribution of transported air pollution to
visibility impairment by the modeling work groups.
2. Further work on the value of visibility in the Eastern North
American context.
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Each case may actually be representative of a variety of aerosols.
For example, case 2 closely approximates the aerosol observed by
Ferman et al. (1981) during their month long study in the Blue Ridge •
Mountains, even though typical .daytime humidities were less than 70%. |
In this study corrected y = 5.5 m /g as measured by heated
nephelometer and when the measured effect of condensed water is •
added, y increases. Thus, even though 90% RH is comparatively rare •
during the daylight hours, case 2 is likely to be closer to typical
summertime eastern conditions than is case 1.
Figure 5-5 shows the powerful effect of humidity and carbon •
absorption on visual range for a given particle level. The curves
also indicate that visibility becomes more sensitive to changes to ]•
fine particle levels below about 100-200 yg/m. Results from this j|
figure should not be compared directly with airport visibility data.
Due to non-ideal targets and observation conditions, airport «
visibilities will tend to be lower than predicted by the Koschmieder •
relationship with K = 3.9. •
When background fine particle concentrations are understood, the ft
Koschmieder equation can be used to relate predicted sulphate levels |
to resulting visual range. Available nonurban summertime fine
particle data are summarized in Figure 5-6. Actual impacts must be •
derived from the results of regional modelling runs provided by •
Workgroup II. If the results of the model are to be "tuned" to
airport visibility data, K in the Koschmieder equation should be
2.9 - 3.5 and y for eastern conditions should be 6-8 m /g. •
5.2.4 Sensitive Areas and Populations •
Clean areas such as found in western North America, are the most
sensitive to visibility degradation. In the U.S., the Clean Air Act «
affords special protection to visibility in 156 'Class I' areas, •
including national parks and wilderness areas. Many of these Class I
areas are located near the U.S./Canada border and one (Roosevelt-
Campobello) in Canada. However, any area, urban or rural, with •
normal viewing distances of a mile or more may be affected by •
episodic regional haze, carrying acid precursor substances.
5.2.5 Data needs/Research Requirements
The following instruments are required to complete work related to •
the effects of atmospheric deposition on visibility: •
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CU
UH
S-i
3
O
c
t)0
3.
0)
Q)
o
•H
§,
0)
c
0>
e
•H
3
CO
bo
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5-36
3. Continued analysis of regional North American visibility data to
further elucidate reliability of data and implications for
anthropogenic contribution.
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5.3 REFERENCES
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pollution health effects report.
Barker, M.L. 1976. Planning for environmental indices: observer
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Barton, S.C., and Johnson, N.D. 1980. Atmospheric deposition of
mercury in Ontario. Report No. P-2699-G04, Ontario Research
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Basu, P.K. 1981. Acid rain and the eye. C.M.A. Journ. 125:338.
Bedi, J.F.; Folinsbee, L.J.; Horvath, S.M; and Ebenstein, R.S. 1979.
Human exposure to sulphur dioxide and ozone: absence of a
synergistic effect. Arch. Environ. Health 34(4): 233-239.
Brookshire, D.S.; D'Arge, R.C.; and Schulze, W.D. 1979. Methods
development for assessing air pollution control benefits.
Vol. II; experiments on valuing non-market goods; A case study
of alternative benefit measures of air pollution control in the
south coast air basin of southern California.
EPA-600/5-79-001b, U.S. Environmental Protection Agency,
Washington, DC.
Brookshire, D.S.; Ives, B.C.; and Schulze, W.D. 1976. The valuation
of aesthetic preferences. J. Environ. Manage. 3:325-346.
Brosset, C., and Svedung, I. 1977. Swedish water and air pollution
laboratory. Report Number B378, Lothenburg and Stockholm.
Brouzes, R.J.P.; McCloar, R.A.N.; and Tomlinson, G.H. 1977. Mercury
- the link between pH of natural water and the mercury content
in fish. Research Report from Domtar Ltd., Research Center,
Montreal, P.Q.
Chang, L.W. 1977. Neurotoxic effects of mercury - a review.
Environ. Res. 14:329-373.
Charlson, R.J.; Waggoner, A.P.; and Thielke, J.F. 1978. Visibility
protection for class I areas; the technical basis. Council on
Environmental Quality, Washinton, DC.
Christe, A. Personal communication.
Cooper, J.A., and Watson, J.G. 1979. Portland aerosol
characterization study. Presented at the 1979 Annual Meeting,
APCA Paper 79-29.4. Air Pollut. Control Assoc., Cincinnati,
OH., 1979.
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1980b. Air traffic service performance measurement
Control Assoc. 29:482-497.
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Covert, D.S.; Waggoner, A.P.; Weiss, R.E.; Ahlquist, N.C.; and
Charlson, R.J. 1980. Atmospheric aerosols, humidity, and
visibility. In Character and origins of smog aerosols. Adv. •
Environ. Sci. Technol. 9:559-581. •
Cronan, C.S., and Schofield, C.L. 1979. Aluminum leaching response
acid precipitation: effects on high-elevation watersheds in the
northeast. Science 204:304-306.
Department of Health, Education and Welfare (DREW). 1969. Air •
quality criteria for particulate matter. U.S. Government
Printing Office, Washington, DC. AP-49.
Doull, J.; Klaassen, C.D.; and Amdur, M.O. 1980. Casarett and I
Doull's Toxicology. New York: McMillan Publishing Co., Inc.
Environment Canada. 1979. National air pollution surveillance - •
annual report. EPS 5-AP-80-15, Air Pollution Control
Directorate, Environmental Protection Service, Ottawa, Ont. ^
Farrell, B.P.; Kerr, T.J.; Kulle, L.R.; Sauder, L.R; and Young, J.L. "
1979. Adaptation in human subjects to the effects of inhaled
ozone after repeated exposure. Am. Rev. Respir. Pis. •
119:725-730. •
Faxvog, F.R., and Roessler, D.M. 1978. Carbon aerosol visibility vs. m
particle size distribution. Appl. Opt. 17:2612-2616. I
Federal Aviation Administration (FAA). 1978. Airmen's information
manual. Basic Flight Information - ATC Procedures. •
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system for major airports, 1975-1980. Washington, DC.
Ferman, M.A.; Wolff, G.T.; and Kelly, N.A. 1981. The nature and •
sources of haze in the Shenandoah Valley/Blue Ridge Mountains ™
area. J. Air Pollut. Control Assoc. 31:1074-1082.
Ferris, B.C. 1978. Health effects of exposure to low levels of •
regulated air pollutants. A critical review. J. Air Pollut.
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Flachsbart, P.G., and Phillips, S. 1980. An index and model of human
response to air quality. J. Air Pollut. Control Assoc. _
30:759-768. •
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Florey, C.; Melia, R.; Chinn, S.; Goldstein, B.D.; Brooks, A.G.;
John, H.; Craighthead, I.B.; and Webster, X. 1979. The relation
between respiratory illness in primary school children and the
use of gas for cooking. III. Nitrogen dioxide, respiratory
illness and lung infection. Int. J. Epid. 8(4):347-353.
Folinsbee, L.J.; Bedi, J.F.; and Horvath, S.M. 1980. Respiratory
responses in humans repeatedly exposed to low concentrations of
ozone. Am. Rev. Respir. Pis. 121:431-439.
Fox, D.; Loomis, R.J.; and Green, T.C. 1979. In Proc. of the
workshop in visibility values, USDA Forest Service, Ft. Collins,
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Freeman, A.M., III. 1979a. The benefits of environmental
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Friedlander, S.K. 1977. Smoke, dust and haze: fundamentals of
aerosol behavior. New York: John Wiley and Sons.
Fuhs, G.W., and Olsen, R.A. 1979. Acid precipitation effects on
drinking water in the Adirondack Mountains of New York State.
State of N.Y. Dept. of Health Tech. Memo., Albany, NY.
Galloway, J.N., and Whelpdale, D.M. 1980. An atmospheric sulphur
budget for eastern North America. Atmos. Environ. 14:409-417.
Goldsmith, J.R., and Nadel, J.A. 1969. Experimental exposure of
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Goldstein, B.D. 1980. Experimental and clinical problems of effects
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Gorham, E. 1958. Bronchitis and the acidity of urban precipitation.
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Grant, W.M. 1974. Toxicology of the eye. Springfield, IL.:
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Groblicki, P.J.; Wolff, G.T.; and Countess, R.J. 1980. Visibility
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Guidotti, T.H. 1978. The higher oxides of nitrogen .'inhalation
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Husar, R.B., and White, W.H. 1976. On the color of L.A. smog.
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Hackney, J.D.; Linn, W.S.; Mohler, J.G.; and Collier, C.R. 1977.
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exposed repeatedly. J. Appl. Physiol. 43(l):82-85. •
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evidence. J.Epid. 110(5):530-659. |
Hultberg, H. , and Wenblad, A. 1980. Acid groundwater in southwestern •
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SNSF-Project, Sandefjord, Norway, 1980.
Husar, R.B. Personal communication. Washington University, St. B
Louis, MO.
Husar, R.B., and Holloway, J.M. 1981. Visibility Trend at Blue Hill, |
Massachusetts, since 1889. Bull. Am. Meteorol. Soc. (in press)
Husar, R.B.; Patterson, D.E.; Holloway, J.M.; Wilson, W.E.; and •
Ellestad, T.B. 1980. Trends of eastern U.S. haziness since ™
1948. In Proc. Fourth Symp. on Atmospheric Turbulence,
Diffusion and Air Pollution, pp. 249-256. American I
Meteorological Society, Reno, NV.
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Jenson, S. , and Jernelov, A. 1972. In Mercury contamination in man _
and his environment, pp. 43-48. Tech. Rep. Ser. 127, IAEA, •
Vienna. ™
Kerr, H.D.; Kulle, J.J.; Mcllhaney, M.L.; and Smipersky, P. 1975. •
Effects of ozone on pulmonary function in normal subjects. An V
environmental chamber study. Am. Rev. Respir. Pis. 111:763-773.
Kleinman, M.T.; Bailey, R.M.; Chang, Y.C.; and Clark, K.W. 1981. |
Exposures of human volunteers to a controlled atmosphere mixture
of ozone, sulphur dioxide and sulphuric acid. Am. Ind. Hyg. _
Assoc. J. 42(l):61-69. •
Latimer, D.A.; Bergstrom, R.W.; Hayes, S.R.; Liu, M.K.; Seinfeld,
J.H.; Whitten, G.Z.; Wojcik, M.A.; and Hillyer, M.J. 1978. The •
development of mathematical models for the prediction of |
anthropogenic visibility impairment. EPA-450/3-78-110a, U.S.
Environmental Protection Agency, Research Triangle Park, NC. «
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Lave, L.B., and Seskin, E.P. 1973. An analysis of the association
between U.S. mortality and air pollution. J. Am. Stat. Assoc.
68(342):284-290.
Leonard, E.M.; Williams, M.D.; and Mutschlecner, J.P. 1977. The
visibility issue in the Rocky Mountain West. Prepared by Los
Alamos Scientific Laboratory for the Dept. of Energy,
preliminary draft report.
Lindberg, S.E. 1980. Mercury partitioning in a power plant plume and
its influence on atmospheric removal mechanisms. Atmos.
Environ. 14:227-231.
Macias, E.S.; Zwicker, J.O.; Ouimette, J.R.; Bering, S.V.;
Friedlander, S.K.; Cahill, T.A.; Kuhlmey, G.A.; and Richard,
L.W. 1980. Regional haze in the southwestern U.S.: I. Aerosol
chemical composition. Presented at the Symposium on Plumes and
Visibility, Grand Canyon, AZ.
Marians, M., and Trijonis, J. 1979. Empirical studies of the
relationship between emissions and visibility in the southwest.
EPA-450/5-79-009, U.S. Environmental Protection Agency, Research
Triangle Park, NC.
Mayfield, 1980. Methylation rates in sediments. M.Sc. Thesis,
University of Windsor, Windsor, Ont.
Menlo, O.T. 1980. Presented at the Symposium on Plumes and
Visibility. Grand Canyon, AZ.
Meranger, J.C., and Khan, T.H. 1982. The impact of lake acidity on
the quality of pumped cottage water in northern Ontario.
Environ. Sci. Technol. (submitted)
Middleton, W.F.K. 1952. Vision through the atmosphere. Toronto,
Ont.: University of Toronto Press.
National Academy of Sciences (NAS). 1977a. Medical and biological
effects of environmental pollutants: nitrogen oxides.
Washington, DC.
. 1977b. Airborne particles. Prepared for Health
Effects Research Laboratory, Research Triangle Park, NC. 554
pp.
. 1977c. Drinking water and health. National
Academy of Sciences, Washington, DC.
. 1978. Sulfur oxides. National Academy of
Sciences, Washington, DC.
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Branch, Health and Welfare Canada, Ottawa, Ont.
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National Health and Welfare (NHW). 1980. Guidelines for Canadian
drinking water quality 1978. National Health and Welfare,
Cat. No. H48-10/1978 - IE, Ottawa, Ont. •
National Park Service (NFS). 1981. National parks statistical
abstract. National Park Service Statistical Service Center, •
U.S. Department of the Interior. •
National Research Council (NRG). 1975. Photochemical air pollution:
formation, transport and effects. NRC Number 14096, Washington, •
DC. •
_ . 1978. Mercury in the environment . National •
Academy of Sciences, Washington, DC. •
National Research Council of Canada (NRCC). 1979. Effects of mercury
_
•
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in the Canadian environment. NRCC Mp§ 16739, Ottawa, Ont.
Patterson, R.K., and Wagman, J. 1977. Mass and composition of an
urban aerosol as a function of particle size for several 9
visibility levels. J. Aerosol Sci. 8:269-279. |
Philips, G.R.; Lenhart, T.E.; and Gregory, R.W. 1980. Relation mm
between trophic position and mercury accumulation among fishes •
from the Tongue River Reservoir, Montana. Environ. Res.
22:73-80
Piersen, W.R.; Brachaczek, W.W.; Truex, T.J.; Butler, J.W.; and ™
Kormiski, T.J. 1980. Ambient sulfate measurements on Allegheny
mountain and the question of atmospheric sulfate in the
northeastern United States. In Aerosols; anthropogenic and
natural, sources and transport, eds. T.J. Kneip and P.J. Lioy.
Ann. N.Y. Acad. Sci. 338:145-173. _
Randall, A.; Ives, C.; and Eastman, C. 1974. Bidding games for
valuation of aesthetic environmental improvements. J. Environ.
Econ. Manage. 1:132-149. I
Rowe, R.D.; d'Arge, R.C.; and Brookshire, D.S. 1980. An experiment
on the economic value of visibility. J. Environ. Econ. Manage. O|
7:1-19. |
Rowe, R.D., and Chestnut, L.G. 1981. Visibility benefits assessment .
guidebook. EPA-450/5-81-001, U.S. Environmental Protection •
Agency, Research Triangle Park, NC. *
Rudy, J. 1980. The McGill methylmercury study. Medical Services M
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5-43
Scheider, W.A.; Jeffries, D.S.; and Dillon, P.J. 1979. Effects of
acidic precipitation on Precambrian freshwaters in southern
Ontario. J. Great Lakes Res. 5:45-51.
Schindler, D.W. 1980. Experimental acidification of a whole lake.
In Proc. Int. Conf. Ecological Impact of Acidic Precipitation,
eds. D. Drablos and A. Tollan, pp. 370-374. SNSF - Project,
Sandefjord, Norway, 1980.
Schusky, J. 1966. Public awareness and concern with air pollution in
the St. Louis Metropolitan Area. J. Air Pollut. Control Assoc.
16:72-76.
Schwenk, J.C. 1981. General aviation and avionics survey. Annual
Summary Report, 1979 Data, FAA-MS-81-1, Transportation Systems
Center, U.S. Department of Transportation.
Sharpe, W.E.; DeWalle, D.R.; and Izbiki, J. 1980. Presented at the
Int. Conf. Ecological Impact of Acid Precipitation.
SNSF-Project, Sandefjord, Norway, 1980.
Sloan, C.S. 1980. Visibility trends II: Mideastern United States
1948-1978. GMR-3474, ENV #92, Environmental Sciences Dept.,
General Motors Research Laboratory Report, Warren, Ml.
Snelling, R. Personal communication. Data from Western Fine Particle
Data Base, Environmental Monitoring Systems Laboratory, U.S.
Environmental Protection Agency, Las Vegas, NV.
Speizer, F.E.; Ferris, B.; Bishop, Y.M.; and Spengler, J. 1980.
Respiratory disease rates and pulmonary function in children
associated with N0£ exposures. Am. Rev. Respir. Pis.
121:3-10.
Stevens, R.K.; Dzubay, T.G.; Russworm, G.; and Rickel, D. 1978.
Sampling and analysis of atmospheric sulfates and related
species. Atmos. Environ. 12:55-68.
Stevens, R.K.; Dzubay, T.G.; Shaw, R.W., Jr.; McClenny, W.A.; Lewis,
C.W.; and Wilson, W.E. 1980. Characterization of the aerosol in
the Great Smoky Mountains. Environ. Sci. Technol.
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quality and morphometric parameters on mercury uptake by
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Ont.
Tang, I.N. 1980. Deliquescence properties and particle size charge
of hydroscopic aerosols. In Generation of aerosols and exposure
facilities for exposure experiments, ed. K. Willeke, pp.
153-167. Ann Arbor, MI.:Ann Arbor Science.
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Taylor, F.B. 1982. A cooperative study of the effects of acid rain
on water supplies. Presented at the Environmental Protection
Agency Peer Review on Acidic Deposition. North Carolina State
University, Raleigh, NC.
Park, NC.
1979. Protecting visibility; An EPA report to
1980. Air quality criteria for oxides of nitrogen.
. 1981. Air quality criteria for particulate matter
and sulfur oxides. Environmental Criteria and Assessment
Office, Research Triangle Park, NC.
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Tomlinson, G.H. 1978. Acidic precipitation and mercury in Canadian «
lakes and fish. Presented at Public Meeting on Acid •
Precipitation. The Committee on Environmental Conservation, New *
York State Assembly and the Adirondack Park Agency, Lake Placid,
NY. •
Tomlinson, G.H.; Brouzes, R.J.P.; McLean, R.A.N.; and Kadlecek, J.
1979. Domtar Research Centre, Senneville, P.Q. •
Trijonis, J., and Shapland, R. 1979. Existing visibility levels in
the U.S.: isopleth maps of visibility in suburban/nonurban
areas during 1974-1976.EPA-450/5-79-010, U.S. Environmental •
Protection Agency, Research Triangle Park, NC. ™
Trijonis, J.; Yan, K.; and Husar, R.B. 1978a. Visibility in the
southwest: an exploration of the historical data base.
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Environmental Protection Agency, Research Triangle Park, NC. «
. Visibility in the northeast: long term visibility
trends and visibility/pollutant relationships.
EPA-600/3-78-075, U.S. Environmental Protection Agency, Research I
Triangle Park, NC. •
Turk, J., and Peters, N.E. 1978. Presented at Public Meeting on Acid •
Precipitation. The Committee on Environmental Conservation, New |
York State Assembly and the Adirondack Park Agency, Lake Placid,
NY. ^
U.S. Environmental Protection Agency (USEPA). 1978. Air quality ~
criteria for ozone and other photochemical oxidants.
Environmental Criteria and Assessment Office, Research Triangle •
Congress. EPA-450/5-79-008, Office of Air Quality Planning and •
Standards, Research Triangle Park, NC.
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Environmental Criteria and Assessment Office, Research Triangle
Park, NC.
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. 1982. Review of the national ambient air quality
standards for particulate matter; assessment of scientific and
technical information. OAQPS Staff Paper, EPA-450/5-82-001,
Research Triangle Park, NC.
U.S. Senate. 1963. Committee on public works, a study of pollution -
air. U.S. Congress, Washington, DC. p. 21.
Waggoner, A.P.; Weiss, R.E.; Ahlquist, N.C.; Covert, D.S.; and
Charleson, R.J. 1980. Optical characteristics of atomsopheric
aerosols. Presented at the Symposium on Plumes and Visibility
Measurements and Model Components. Grand Canyon, AZ.
Wall, G. 1973. Public response to air pollution in South Yorkshire,
England. Environment and Behaviour 5:219-248.
Ware, J.; Thibodeau, L.A.; Speizer, F.E.; Colone, S.; and Ferris,
E.G. 1981. Assessments of the health effects of sulphur oxide
and particulate matter: analysis of the exposure - response
relationship. Environ. Health Perspect. (in press)
Watson, J.; Chow, J.C.; and Shah, J.J. 1981. Analysis of inhalable
and fine particulate matter measurements.EPA Contract #
68-02-2542, Task #6, Final report December 1981.
Weiss, R.E.; Waggoner, A.P.; and Charlson, R.J. 1978. Studies of the
optical, physical, and chemical properties of light-absorbing
aerosols. In Proc. Conf. on Carbonaceous Particles in the
Atmosphere, pp. 257-262.
Wolff, G.T.; Groblicki, P.J.; Cadle, S.H.; and Countess, R.J. 1980.
Particulate carbon at various locations in the United States.
Presented at the General Motors Symposium on Particulate Carbon.
Warren, Ml.
World Health Organization (WHO). 1979. Environmental health criteria
(8): sulfur oxides and suspended particulate matter. WHO,
Geneva, 1979.
Wren, C.; MacCrimmon, H.; Frank, R; and Suda, P. 1980. Total and
methyl mercury levels in wild mammals from the Precambrian
shield area of southcentral Ontario, Canada. Bull. Environ.
Contam. Toxicol. 25:100-105.
Young, W.A.; Shaw, D.B.; and Bates, D.V. 1964. Effect of low
concentrations of ozone on pulmonary function in man. J. Appl.
Physiol. 19:765-768.
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SECTION 6
EFFECTS ON MAN MADE STRUCTURES
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6-1
SECTION 6
EFFECTS ON MAN-MADE STRUCTURES
6.1 INTRODUCTION
Previous sections of this report have focused primarily upon the
effects of pollutants on sensitive receptors in remote wilderness
areas. In this section, the receptors (i.e., man-made structures)
are usually co-located with the pollution sources. The distinction
between the effects of pollutants from near or intermediate sources
(i.e., 10's to perhaps 100 km away) and from distant sources (i.e.,
100's to perhaps 1000's of kilometres away) is difficult if not
impossible to make. This is particularly the case when local primary
species, S02 for example, oxidize to secondary products (SO/p")
at a nearby site which is also receiving SO^- from distant sources
of primary S02 which has oxidized during its atmospheric residence
time. In most cases, the atmospheric load from local sources tends
to dominate the low concentrations arriving from remote sources
upwind.
In the context of material deterioration, the distinction between
local and distant sources may be academic since damage in general
would be reduced through a reduction in concentration of the major
agents, regardless of the distance these pollutants traveled to the
deposition site.
Consideration of damage will be limited to exterior surfaces, not
only because the wet deposition of pollutants and surface wetting is
primarily limited to exterior surfaces, but also because the
concentrations of corrosive species are usually much higher outside
than inside buildings. Textiles and fabrics are usually associated
with confined environments and are beyond the scope of this section.
Although the economic consequences of material deterioration due to
air pollution are discussed elsewhere in this report (see Section
8.5), this section contains a brief description of some of the
attempts which have been made to quantify damage in economic terms
and the inadequacies of these assessments. For sulphur dioxide,
perhaps the most important corrosive agent, direct costs of
duplication, replacement or protection of certain materials can be
approximated. Before the economic implications can be adequately
addressed, a better understanding is needed of the dose-response
relationships of materials to different pollutants, of the
distribution of materials, and of the replacement and maintenance
factors.
In this section, there is discussion of the effects of four types of
pollutants; S02, N0£ and 03, NH3, and particulates on four classes of
materials; metals and alloys, coatings, masonry, and elastomers.
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There are several quantitative examples of calculated erosion and
corrosion of materials over time.
6.2 OVERVIEW
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All building materials degrade to some extent with time, even in the |
absence of air pollution. Hence, it is important to differentiate
between expected weathering, and accelerated deterioration •
attributable to air pollutants. There are reviews available of the •
weather factors affecting the corrosion of metals (Ashton and Sereda
1981; Sereda 1974) and the durability of stone (Julien 1884),
concrete (Maslow 1974), and polymers (Eurin 1981). In northern •
climates, deterioration effects due to atmospheric pollution may be •
masked in winter by the impact of road deicing salts (CaCl2 and/or
NaCl), which damage porous masonry and are particularly corrosive to •
metals. As noted above, the assessment of deterioration, due to f
pollutants transported over long distances, is confounded by the
impact of pollution produced locally. _
The apparent chemical/physical degradation processes resulting from *
interactions of materials with pollutants and naturally occurring
atmospheric constituents have been reviewed in the literature B
(Benarie 1980; USEPA 1978, 1981a,b; Yocom et al. 1982). A number of V
field studies estimate the relationship between deterioration and
atmospheric deposition without full environmental characterization of •
the test sites, including measurement of meteorological and air •
quality variables. Laboratory tests have also been used to quantify
materials damage from pollutants. However, quantitative
relationships derived from chamber tests cannot be used directly to •
predict damage to exposed materials. ™
At the outset it must be acknowledged that the objective assessment •
of the response of materials to corrosive agents cannot provide |
adequate estimates of "loss" resulting from deterioration of
historic materials. Monuments must be distinguished from other •
structures because here the net loss by deterioration embraces I
aesthetic and historical contributions where monetary scales may not
apply. Trade-offs made in mitigating the impact of air pollutants
should address the preservation of the qualities that constitute the •
significance of the monument. ™
The architectural and sculptural expressions of our two heritages are •
a precious nonrenewable resource. Historic structures and monumental J[
statuary represent the most visible aspects of historical and
cultural evolution. In the United States, legislative recognition of •
the value of this cultural heritage, giving a mandate for its •
preservation, began in 1906, with the passage of the Antiquities Act,
and continues to this day, with the passage of the Historic
Preservation Act Amendments of 1980. The 1916 Organic legislation of •
the National Park Service gives a mandate for the conservation of •
"...'historic objects' .. .to provide for the enjoyment of the same in
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6-3
such a manner and by such means as will leave them unimpaired for the
enjoyment of future generations."
In Canada, there is similar legislation. For example, the
Archeological Sites Protection Act was enacted in 1953, the same year
the Historic Sites and Monuments Act was adopted.
Studies of the economics of damage due to air pollution are contained
in papers by Haynie (1980a), Yocom and Upham (1977), Spence and
Haynie (1972), Waddell (1974), Liu and Yu (1976), Yocom et al.
(1981), and Yocom et al. (1982) among others. In these instances,
dose-response relationships were obtained by relating the
concentrations present in the atmosphere to the degree of damage
observed or measured. As an example, a report released by the
Organization for Economic Cooperation and Development (OECD 1981),
indicates that for the eleven European countries studied, based on an
expected production of 24.4 million metric tons of SC>2 in 1985, the
resulting benefit of a 50% reduction in 862 emissions would be of
the order of $1.16 billion per year in terms of reduced damage. In
this case, the benefit is based on corrosion of a limited number of
metals and excludes masonry and coatings since no exact dose-response
relationships are available between deposition of sulphur compounds
on these materials and deterioration rates. The loss due to the
impoverishment of cultural heritage is not assessed. As noted
earlier, the more difficult task of relating this damage to human
response (i.e., replacement, substitution or protection) has not yet
been completed and will be more fully discussed in subsection 6.8.
6.3 MECHANISMS AND ASSESSMENT OF EFFECTS
The quantitative expression of a relationship between exposure to a
particular pollutant, and the type and extent of the associated
damage to a specific material is known as a dose-response relation-
ship or a damage function (Hershaft 1976). The damage function
should express a quantitative cause-to-effect link (Benarie 1980).
A significant mathematical correlation between a measured damage and
an ambient pollutant concentration is not sufficient proof of a
causal relationship. Because a particular kind of damage may also
occur in the absence of the pollutant(s), the damage function should
explicitly describe the incremental effect attributable to the
pollutant(s). Moreover, in some cases for metals, a certain minimum
dose or threshold value is required before an effect is observable
and at high doses a saturation level may also be observed.
6.3.1 Factors Influencing Deposition
The relation between concentration of pollutants in ambient air and
in precipitation and the amount of pollutant delivered to a
material surface is known as the deposition velocity. Deposition is
a two-step phenomenon beginning with delivery of the pollutant to the
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surface controlled by aerodynamics and followed by pollutant -
material interaction. Gaseous deposition velocities can be as rapid
as 0.021 - 0.6 m/min (Judeikis 1979). However, large variations are •
found in measured deposition velocities (McMahon and Denison 1979). |
This scatter is in part due to intrinsic properties of materials and
in part due to extrinsic factors (e.g., surface moisture, wind speed •
and temperature). •
Acidic deposition can occur on materials under both wet and dry
conditions. Under wet conditions, the aqueous form of the pollutant, I
(e.g., sulphuric acid) reacts with the material in question to form •
reaction products. These products may be either more or less soluble
than the original material. The detailed kinetics of these reactions •
are not well known and variations in reactions may occur due to the •
surface conditions of the material and the presence of additional
chemical species. _
Deposition may also occur under dry conditions. In theory, a gaseous •
pollutant such as S02 or NOx can react directly with metals and
masonry without going through an aqueous phase (Torraca 1981a; Van •
Houte et al. 1981). As a practical matter, the reactions in the |
environment always involve the presence of humidity. Consequently,
reaction rates have not been studied under completely dry conditions •
(Haynie and Upham 1974; Judeikis and Stewart 1976; Spedding 1969). •
It is assumed that dry deposition reactions do in fact involve an
intermediate stage where the gaseous pollutants are oxidized in the
presence of available surface moisture and proceed to attack the •
materials in aqueous form. The major distinction between this kind •
of deposition and what is termed wet deposition is that in the latter
the moisture involved comes only from precipitation. Dry deposition •
involves all sources of moisture other than precipitation. f
These other sources of moisture include surface condensation, which _
occurs under certain conditions of relative humidity, dew point, and •
surface temperature. There may also be internal condensation arising •
when warm, humidified indoor air attains the dew point within a
cooler masonry wall (ICOMOS 1967; Torraca 1981a; Winkler 1975). •
Groundwater wicked up through masonry walls by capillary action is |
another source for moisture in masonry walls (Melville and Gordon
1973). Porous materials (e.g., stone, brick and concrete) may retain •
significant amounts of moisture in the pores even during prolonged •
spells of dry weather.
Surface moisture has a strong influence on deposition velocities. •
It has also been suggested that the rate-limiting step for deposition ™
of pollutants may be the atmospheric transport processes that
controls the delivery of gases through the quasi-laminar boundary at •
the surface (Hicks 1981). These transport processes would increase |
with increased local wind speed (Lawrence 1962).
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6-5
Particulate matter on the surface can also affect the deposition
velocity. Hygroscopic particulates like marine salt tend to
increase the time the surface is wet (Fassina 1978). Elements
commonly found in particulates (e.g., iron, manganese, and copper)
have been identified as serving as catalysts in the oxidation of
S02 to S042~ (Hegg and Hobbs 1978).
The condition of the surface itself will also influence the
deposition velocity. A material that has been exposed for a
significant amount of time may show a deposition velocity different
from the initial one (Judeikis 1979). In the case of most metals,
iron being the notable exception, a tight layer of reaction products
(e.g., oxides, carbonates and chlorides) will form on the surface
giving some degree of protection. In stone, however, the coating of
reaction products remains porous and the rate of attack does not
decrease (Van Houte et al. 1981). In fact, it may increase as the
effective surface increases (Judeikis and Stewart 1976) or as the
greater roughness increases the atmospheric transport processes
(Hicks 1981).
Exposure specifications such as the siting of a structure or a
material's position in a structure must also be considered. The most
important condition is the exposure of the material to rain. In
addition to providing an aqueous medium for acidic attack, the runoff
of rain water serves as a major agent in the dissolution and removal
of weathering products. For example, it has been observed on the
Acropolis in Athens and at the Field Museum in Chicago that the
marble has reacted with S(>2 to form an even layer of calcium
sulphate. This layer several centimetres thick remains intact on
some surfaces that are not washed by rain while on other parts of the
structure washed by rain, the layer does not exist (Gauri 1979;
Skoulikidis et al. 1976).
The rate of deposition will increase with increased wind speed.
Although primarily determined by the prevailing climate at the
location, the wind speed at a given point on a surface can be
influenced by the orientation of the structure, architectural
details, and by other buildings in the vicinity (BRS 1970; Kotake and
Sano 1981; Lacy 1971). For example, wind speeds around a building
will vary with height and increase at corners and over cornices (Lacy
1971). The corrosion rates for galvanized wire and fencing are
almost double that of galvanized sheet, indicating the influence of
the geometry of a material on wind speed (Haynie 1980b). Erosion of
materials by wind-borne abrasive particulates may also be significant
at some locations. Finally, wind will cause rain to fall on a slant
rather than perpendicularly to the ground. Thus it will drive
rainfall onto vertical surfaces that otherwise would remain dry
(Marsh 1977).
Finally, the conditions of thermal exposure will vary around a
structure. This occurs depending on compass orientation, angle from
vertical, and shading. Each part of the structure will receive
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6-6
6.3.2 Effect of Sulphur Dioxide Pollutant/Material Interactions
H2S04 + Zn 5=^ ZnS04 + H£(gas)
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differing amounts of solar radiation, thus will show differing
ranges in diurnal temperature cycles. Temperature cycles are also a
function of each material's thermal response characteristics. •
Man-made sources of heat, either within a building or from nearby "
sources will also modify the local thermal environment. Ranges in
thermal conditions influence the atmospheric transport processes, as
well as moisture content of material surfaces and thus, deposition
velocities.
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There is general agreement in the literature that S0£ is the B
primary species causing damage to materials exposed to the •
atmosphere (Boyd and Fink 1974; Barker et al. 1980; Kucera 1976;
Mansfeld 1980; Mikhailovskii and Sanko 1979; OECD 1981; Yocom and •
Upham 1977). |
6.3.2.1 Zinc
The specific S0£ reactions that cause metal corrosion are not •
fully understood. Some of the possible mechanisms involved have been
discussed (Benarie 1980; USEPA 1981a). For the relatively simple •
case of zinc, the overall reaction is B
S02 + Q£ + Zn ^ ZnS04 «
Since ZnSC>4 is soluble and readily lost from surfaces exposed to
rain, a protective surface film is not formed. In cases where the
S02 is converted to acidic sulphate prior to reaction with the •
zinc, the expected reaction would also result in the same soluble B
zinc species
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Since the corrosion rate for zinc is high for solutions having a pH
less than about 5 (Pourbaix 1966), many acidic species are expected B
to cause the corrosion of zinc. For the case of precipitation, it ™
should be noted that the annual average pH for most of eastern North
America is below 5 (see Figure 2-4). B
In a recent publication (Haynie 1980b), data from six different
studies for the atmospheric corrosion of zinc were reevaluated with •
respect to the following relationship: B
Cz = ATW + BTWS02 (1) _
where Cz = zinc corrosion in micrometres ™
TW = time of wetness in years
S02 = average concentration of SC>2 in I'g/m-^ B
A and B = regression coefficients.
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6-7
The results of the coefficients evaluation (Haynie 1980b) are shown
in Table 6-1.
As is typical for many atmospheric corrosion tests, intercomparisons
between studies are subject to many problems due to differences in
test objectives, techniques, and available environmental monitoring.
In the studies examined by Haynie (1980b), the time of wetness was
calculated in one of three ways: (1) by using the average relative
humidity employing a previously evaluated empirical equation
(Cavender et al. 1971; Haynie et al. 1976); (2) by using the total
time the relative humidity exceeded 85% (Haynie et al. 1976) or 90%
(Mansfeld 1980); and (3) by using a "dew-detector" (Guttman 1968;
Guttman and Sereda 1968). The average concentration of S02 was
determined by using continuous instruments (Haynie et al. 1976;
Haynie and Upham 1970; Mansfeld 1980) using lead peroxide techniques
(Cavender et al. 1971; Guttman and Sereda 1968) or both (Guttman
1968). In spite of the experimental differences, it is apparent that
S02 is an important cause of degradation of zinc. The B
coefficient is lower for the chamber study than for the field-
determined values (Haynie et al. 1976). While this difference was
attributed (Haynie 1980b) to a lower gas velocity in the chamber
compared to the field studies, it has been suggested (Yocom et al.
1982) that the higher values reported for the field studies may be
the result of the combined effect of S02 and particulate matter
containing sulphates, chlorides, nitrates and other anions. The low
B coefficient determined in the St. Louis study (Mansfeld 1980) may
also indicate that lower particulate levels result in lower S(>2
corrosion rates. The tacit assumption being made here is that during
the test period the particulate levels in St. Louis were low.
Haynie (1980b,c) has demonstrated that the average wind speed has an
important influence on the S02 deposition velocity and, hence, the
corresponding corrosion rates for zinc. Also, the deposition
velocity has also been shown to be dependent on geometry. For
example, the corrosion rate for galvanized wire and fencing has been
shown (Haynie 1980b) to be approximately twice that of galvanized
sheet exposed to the same environment. Since the lead peroxide
method was used to determine total sulphur (as SC>2) in two of the
studies (i.e., Cavender et al. 1971 and Guttman and Sereda 1968 as
shown in Table 6-1), those values would be expected to be
wind-velocity dependent (Lynch et al. 1978). In contrast, the SC>2
values measured by continuous instruments should be independent of
wind velocity. The possibility that this wind-dependence affected
the SC>2 coefficients should be considered before such data are used
for any damage function calculations.
6.3.2.2 Steels
A number of atmospheric exposure studies have been conducted for
steels (see reviews cited at beginning of Section 6.2). Only a few
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6-8
Table 6-1. EXPERIMENTAL REGRESSION COEFFICIENTS WITH ESTIMATED
STANDARD DEVIATIONS FOR SMALL ZINC AND GALVANIZED STEEL
SPECIMENS OBTAINED FROM SIX EXPOSURE STUDIES
Haynie and Upham (1970) 1.15+0.60 0.081+0.005 37
Mansfeld (1980) 2.36 + 0.13 0.022 + 0.004 156
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A B Number of
Reference ym/year ym/(g/m3)year Data Sets •
Cavender et al. (1971) 1.05 + 0.96 0.073 + 0.007 173 •
Guttman (1968) 1.79 0.024 large
Guttman and Sereda (1968) 2.47 _+ 0.86 0.027 _+ 0.008 136 I
Haynie et al. (1976) 1.53 + 0.39 0.018 +_ 0.002 96
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6-9
studies have been conducted where simultaneous air quality and
deposition measurements have been made. In general the corrosion
rate for uncoated low-carbon steels has been found to depend on
S02 concentration and exposure time . For example , in one study
(Haynie and Upham 1971) the following relationship was developed
Cs = 9.013 [eO.00161 S02] [ (4. 768t)0-7512 - 0.00582 OX] (2)
where Cs = corrosion in micrometres
t = exposure in years
S02 = average concentration of S02 in y g/m^
OX = total oxidants in yg/m^ (see discussion of
ozone below) .
In this study, similar expressions were developed for a weathering
steel. Although humidity effects were not determined, the humidity
was approximately the same at all the sites used. In another study,
Haynie and Upham (1974) reported a similar functional dependence
Cs = 325 tl/2 exp[0. 00275 S02 - >] (3)
where RH is the relative humidity. In this case, oxidants were not
measured. Equations 2 and 3 both indicate that the corrosion rate
increases with S02 concentration but decreases with exposure time.
Equation 3 predicts very low corrosion rates for low humidity,
regardless of the S(>2 level. However, at high humidity the steel
would corrode even in the absence of S02- Although equations 2 and
3 may give a correct description of the interaction of S02 with low
carbon steels, they have little relevance, since virtually no steels
of this type are unprotected in the environment. However, studies of
this material have proven useful in estimating the relative
corrosivities of various exposure sites. Even in the case where a
painted steel begins to rust, equations 2 and 3 probably do not
accurately describe the corrosion occurring.
For a weathering steel, a material normally left uncoated, Haynie and
Upham (1974) reported a relationship of the form shown in equation 2,
with a lower S02 coefficient and a slightly different time
dependence. On the other hand, studies by Copson (1945), Cramer
et al. (1980), and Suzuki et al. (1980), found that sulphur
incorporated into the film improved the corrosion resistance for a
weathering steel , and that this sulphur could be related to the
average S02 level in the local environment. The amount of copper
and tin in the steel has been shown to influence its corrosion
resistance in sulphur containing atmospheres (Cramer et al . 1980).
6.3.2.3 Copper and Copper Alloys
Copper and its alloys are among the most durable materials for
exterior exposure because they form protective patinas. The pH
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6-10
6.3.2.4 Aluminum
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value of precipitation is probably a significant factor determining
the development of basic carbonates and/or oxide protective patina
(Guttman and Sereda 1968; Mattsson and Holm 1968). Below pH of about I
4, the protective coating may be rapidly dissolved exposing the bare "
metal (Allaino-Rossetti and Marabelli 1976; Gettens 1964; Pourbaix
1966). It is noteworthy that the mean annual pH of precipitation is •
near 4 in portions of northeastern United States, southern Ontario |
and Quebec (Figure 2-4). The presence of S02 in the atmosphere
appears to accelerate the formation of the green patina (Mattsson and _
Holm 1968), normally desired on architectural copper surfaces. •
Analysis of the patina of the Statue of Liberty, for example, ™
indicates that it is predominately copper sulphate, with less than 1%
of copper carbonate and copper chloride (Osborn 1963). Exposure I
tests indicate that the most uniform patina forms on unalloyed •
copper. The presence of alloying elements, primarily tin and zinc
for bronze and brass, respectively, interferes with patina formation •
and thus lowers corrosion resistance of the alloy (Scholes and Jacob •
1970; Walker 1980).
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There are no quantified damage functions reported for aluminum,
although some evidence exists for both S02~induced damage •
(Benarie 1980) and for particulate-assisted S02 pitting damage |
(USEPA 1981a). Sulphur dioxide may also play a role in
stress-assisted corrosion problems for aluminum (Gerhard and Haynie •
1974; Haynie et al. 1976). •
Several relatively new aluminum and aluminum-zinc coated steels have
become available commercially in the past few years. While no •
damage functions have been reported, these coatings have been shown •
to offer superior performance to galvanized steels, with improvement
in service life by a factor of from two to four times in marine, •
rural and industrial environments (Zoccola et al. 1978). These |
coatings will have a great influence on future materials selection
and on estimates of future damage cost related to coated steels. _
6.3.2.5 Paints •
A few investigations, which have been reviewed in several recently I
published documents, have studied the effects of gaseous pollutants m
on the performance of exterior coatings (USEPA 1981a; Yocom et al.
1982). Several of these studies have shown that S02 can penetrate •
into the paint film (Svoboda et al. 1973; Walsh et al. 1977). •
Holbrow (1962) found sulphites and sulphates to be present after the
absorption of S02 by the paint film. In more recent studies, the
high erosion rate observed for oil base house paints was associated •
with the loss of calcium carbonate, an extender pigment of the paint. •
These controlled studies involved the exposure of several types of
paints to atmospheric pollutants under a prescribed dew-light cycle
(Campbell et al. 1974; Spence et al. 1975).
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6-11
A clear mechanism for the deterioration of coatings by wet or dry
pollutant deposition cannot be derived from these investigations.
However, it is apparent that moisture on the painted surfaces plays
an important role by collecting pollutants, especially SC>2, and
thereby forming an acidic aqueous media that facilitates reaction
with the paint film.
The study of Spence et al. (1975) is the only investigation for
which pollutant dose-response relationships have been derived.
Linear regression relationships were developed for two types of
coatings:
Oil Base E = 14.3 + 0.0151 S02 + 0.388 RH (4)
Vinyl Coil E = 2.51 + 1.60-10~5 x RH x S02 (5)
where E = erosion rate in ym/yr
SC>2 = average concentration of SC>2 in yg/m^
RH = relative humidity in percent.
The oil base household paint was found to have a higher erosion rate
which was strongly correlated with the concentration of S02 and
relative humidity. However, the rate is more sensitive to changes in
the humidity than S02. For the vinyl coating, the S02 effect is
statistically significant but contributes less than 5% to the film
erosion rates at ambient levels of concentration. These functions
were obtained under controlled conditions of simulated sunlight and
high temperatures and should not be applied directly to ambient
conditions.
Table 6-2 provides examples of material loss in one year due to a
range of sulphur dioxide concentrations and specific values of the
other variables contained in equations 1 to 5. The table is intended
to provide an indication of corrosion rates for a range of conditions
which might be encountered at sites throughout North America.
6.3.2.6 Elastomers
A chamber study has been conducted in which samples of auto sidewall
tires were exposed to two levels (0.1 and 1.0 ppm) of 03, S02 and
NC>2. The tires were not found to be affected by S02 (Haynie
et al. 1976). Observations of S02 damage to elastomeric materials
have not been reported (USEPA 1981a).
6.3.2.7 Masonry
Masonry materials are porous inorganic substances including stone
and man-made composites such as brick, terra cotta, concrete,
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6-12
TABLE 6-2. EXAMPLES OF MATERIAL LOSS IN ONE YEAR, L IN Urn USING EQUATIONS 1 TO 5 AND
TYPICAL AMBIENT VALUES.
(A)
(B)
(C)
(D)
(E)
Zinc
Steel
Steel
Paint
Paint
(Low carbon)
(Weathering)
(Oil base)
(Vinyl)
L2
Lsl
Lsw
Lpo
Lpv
= 1.79
= 9.013
= 325 +
= 14.3
= 2.51
Tw + 0.024 TwS02
EXPI 0.0016 S02)
i EXPI 0.0027 5 SO
+ 0.0151 S02 + 0.
+ 1.60 10 x RH
[4.768°-7512-0-005820X]
1.63.2 ,
2 RH
388 RH
x S02
(Eq.
(Eq.
(Eq.
(Eq.
1)
3)
4)
5)
S02 Pg/m3 a
10
25
40
55
(A)b
(B)c
(C)d
(D)e
(E)f
Tw = 0.05
= 0.1
= 0.2
OX = 0.05
= 0.50
= 1.00
RH = 40
= 60
= 80
RH = 40
= 60
= 80
RH = 40
= 60
= 80
0.102
0.203
0.406
27.84
27.72
27.60
5.65
22.01
43.44
29.97
37.73
45.49
2.516
2.520
2.523
0.120
0.239
0.478
28.52
28.40
28.27
5.89
22.93
45.27
30.20
37.97
45.72
2.526
2.534
2.542
0.138
0.275
0.550
29.22
29.10
28.96
6.14
23.90
47.17
30.42
38.18
45.94
2.536
2.548
2.561
0.156
0.311
0.622
29.93
29.81
29.67
6.39
24.91
49.16
30.65
38.41
46.17
2.545
2.563
2.580
a S02 concentrations are highly variable within urban areas but normal ly He within the
range 10 to SOyg/ntj depending on city size and industrial activity.
b The coefficients 1.79 and 0.024 are used for Illustration purposes and lie within the
ranges given In Table 6-1. Note the high degree of dependence on wetness as well as
S02 concentration. Wetness values are approximation for south central (0.05), central
(0.1) and coastal (0.2) sites in North America.
c The function depends upon S02 concentrations, as stated In the text above, but appears
to be not strongly dependent on OX. In this example OX Is taken to mean Oj wtth
values given being representative of concentrations In clear air (0.05), smoggy air
(0.5) and episodes (1.00).
d The values cited are for the first year of exposure. Relative Humidity values represent
south central, central and coastal sites. This damage function is strongly dependent on
RH. In addition, material loss is well correlated with S02 concentration. Note the
much lower loss of zinc (A) compared with weathering steel (C).
e The erosion of oil-base paint Is strongly dependent on relative humidity and much less
dependent on S02 concentration.
f As noted in the text, the deterioration of vinyl paints is not strongly dependent on
S02.
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6-13
mortar, stucco, and adobe. Degradation of these materials in the
atmosphere involves disruption of the interlocked mineral components
through chemical and mechanical processes. Mechanical degradation
disrupts the physical structure, through differential thermal
expansion, freeze/thaw, salt crystallization, hydration, and
migration, or by intrusion of root fibriles into the masonry matrix
(Torraca 1981a; Winkler 1978). Chemical processes are those where
ions react with the material and alter the mineral composition to
form weathering products. If less soluble than the original
material, the weathering products will remain as crusts or
discolorations on the masonry surface. Soluble weathering products
will be washed away with sufficient rain water flow, eroding the
material surface. In the absence of runoff, the soluble salts may be
transported into the body of the masonry, and there trigger
mechanical weathering effects such as subflorescence and spalling.
Degradation rates depend not only on the chemical composition of the
mineral composite, but also on grain size and porosity of the matrix
(Jakucs 1977). These factors vary not only between general classes
of masonry materials, but also from quarry to quarry, from concrete
mix to concrete mix. Masonry materials of varying composition, or
varying in grain size and porosity, will not exhibit similar
degradation rates. Additionally, homogeneity of mineral composition
plays a role in the durability of stone building materials.
Inclusions and veins of minor constituents (e.g., as feldspars or
micas) provide zones of preferential weathering creating microcracks
and fissures and exposing interior areas to degradation processes.
As such, any dose-response functions will tend to be material
specific making generalizations difficult.
Commonly used masonry materials are primarily composed of carbonates
or silicates. Silicates are generally more resistant to dissolution
by atmospheric acids than carbonates (Loughnan 1969; Winkler 1975).
Work on the interaction of SOX and masonry material has been
directed towards describing the processes and end products
qualitatively (see Stambolov and van Asperen de Boer 1976 for
references) and towards estimating deposition velocities (Braun and
Wilson 1970; Judeikis 1979; Spedding 1969). Difficulties arise in
determining dose/response relations partly because of the length of
time required to produce measurable effects in both field and chamber
studies (Trudgill 1977) and partly because of imperfect simulations
of real-world cycles of temperature and moisture in chambers
studies.
Of the commonly found construction materials, carbonate minerals have
been studied more extensively than other minerals. The mechanism of
S02 attack on calcite, the major mineral constituent of carbonate
rocks and cementing materials of some sandstones, may proceed through
several mechanisms according to the following equations (Gauri and
Holden 1981; Torraca 1981a):
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6-14
I
Dry conditions: CaC03 + SC>2 + 1/2 02 — > CaS04 + C02
with a possible CaS03 intermediate and •
Wet conditions: CaC03 + H2S04 — > CaS04 + H20 + C02
•
Cycles of available moisture allow for conversion of anhydrite to
the hydrates; gypsum (CaS04 . 2 H20) and bassanite (CaS04 . 1/2
H20). Associated changes in crystal size and pressure are •
significant factors in mechanical decay processes (Arnold 1976; •
Stambolov 1976; Winkler 1975). Similar pressures are exerted by
hydration of sodium and ammonia sulphates (Torraca I981a). The H
presence of these soluble sulphate salts in the subsurface of the |
masonry can cause spalling of the surface under freeze/ thaw
conditions. M
Calcium sulphate is more soluble than calcium carbonate, although a
range of solubilities have been reported for both minerals (Hardie
1967; Jakucs 1977; Keller 1978). In runoff conditions, calcium •
sulphate will dissolve and the material surface will be eroded. •
Calcium sulphate may also combine with other deposits (e.g., carbon
particles and soot) forming black crusts (Fassina et al. 1976; Weaver •
1980). |
Debates concerning the contribution of biological weathering to stone _
deterioration have gone on since the 18th century. Bacteria on the •
surface of buildings convert sulphur dioxide from the atmosphere into ™
sulphuric acid for use as a digestive fluid (Babick and Stotsky 1978;
Winkler 1978). The digestive fluid attacks the calcium carbonate in H
the limstone, marble, or sandstone, liberates carbon dioxide, the •
microbe nutrient, and produces calcium sulphate as a by-product.
Other studies have claimed that all deterioration can be attributed •
to chemical and mechanical processes, and that biological aggression •
is negligible (Fassina 1978; Torraca I981b).
Several special cases of damage to stone structures subjected to •
extremely high local levels of pollution are noted in the M
literature:
moisture and temperature cycles.
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1. The effect of air pollution exposure is illustrated by comparing
the condition of the Elgin Marbles (removed from the Acropolis
at the start of the 19th century) to the sculptures that •>
remained exposed to the atmosphere in Athens. The Elgin I
marbles, kept indoors at the British Museum, are in much better
condition (Skoulikidis et al. 1976), albeit some of the
weathering of the exposed sculptures must be attributed to •
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6-15
2. Some of the blocks in Cologne Cathedral were put in place in the
middle of the 19th century, at the same time that sandstone from
the same quarry was used in building Neuschwanstein Castle in
Bavaria. The Cologne blocks are significantly more deteriorated
than the Neuschwanstein, where the sulphate deposition is lower
by a factor of roughly 20 (Luckat 1976).
3. The Ottawa Parliament House shows considerable deterioration
caused by sodium sulphates trapped in the stone. The source of
the sulphates is presumably a paper mill that operated nearby
until 1973 (Stewart et al. 1981; Weaver 1980).
6.3.3 Effect of Nitrogen Dioxide and Ozone Pollutant/Material
Interactions
6.3.3.1 Metals
There is little information relating NOX concentrations to degrada-
tion rates of metals and masonry. No damage function for NC>2 on
metals is given in USEPA (1981b). Haynie (1980a) assessed corrosion
rates of steels exposed to NOX, S02 and 03 under varying
relative humidity, temperature and wind-speed conditions. While no
dose-response relation has yet been determined for steels, it was
found that oxides of nitrogen (expressed as N02) contributed
significantly to the corrosion response of zinc/copper sensors used
in an "Atmospheric Corrosion Monitor".
A recent study by Byrne and Miller (1980) found that NOX can
influence the corrosion rate of aluminum more than SC^. However,
the author suggests that S02 may reduce nitrogen oxides to nitrogen
gas in the presence of a catalytic surface (e.g., A1203). Hence,
this process may reduce the potential for damage by NOX.
Aqueous nitric acid has a more deleterious effect on most metals,
than H2S04 or HC1 (McLeod and Rogers 1968). In addition, metal
nitrate salts tend to be more soluble than the sulphate salts, so
that nitrate corrosion products can be readily washed from surfaces,
exposing fresh metal to attack. Conversely, sulphate products may
remain on the surface to inhibit further corrosion.
The reported effects of ozone on metal corrosion appear to be
contradictory. While the previously mentioned study by Haynie and
Upham (1971) showed that oxidants correlated with lower steel
corrosion rates (equation 2), a chamber study with ozone showed no
effect on steel corrosion (Haynie et al. 1976). It has been
suggested (Benarie 1980) that the effect observed in the Haynie and
Upham (1971) study was either caused by some oxidant other than ozone
or was related to another factor that was covariant with ozone (e.g.,
temperature or humidity).
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6-16
6.3.3.2 Masonry
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Deposition velocities have been measured for NO and NC>2 on some I
masonry surfaces. The measured deposition velocities for NOX on B
cement appeared to be slightly lower than for SC>2. However,
differences over time of exposure in the behavior of the gases makes •
it impossible to give direct comparison (Judeikis and Wren 1978). |
Since there is no proposed mechanism of ozone effects on masonry
deterioration, nor published studies describing such an effect, _
dose-response relationships are not to be expected. I
6.3.3.3 Paints
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In a chamber study conducted by Spence et al. (1975), ozone was
found to be the most likely factor to affect the erosion rates of
acrylic coil coatings. The following dose-response relationship was •
derived: I
E = 0.159 + 0.000714 03 (6) _
where E = erosion rate in ym/yr •
03 = concentration of 03 in yg/nH _
As indicated by the relationship, the effect of ozone on the erosion ™
rate is negligible, even though it was found to be statistically
significant. I
Vinyl and acrylic coil coatings were not significantly affected by
NC>2 at ambient levels. m
6.3.3.4 Elastomers
The cracking of rubber products results from the combined effects of I
ozone and stress on sensitive elastomeric material. Sensitive Bi
elastomers contain olefin structures (carbon double bonds) which are
susceptible to chemical attack by ozone (Bailey 1958). Natural •
rubber and certain synthetic elastomers (e.g., styrene-butadiene, |
polybutadiene, and polyisoprene) contain these chemical structures.
The ozone-olefinic reaction can result in chain scissioning as well M
as cross-linking of the elastomeric material. In the case of chain I
scissioning, the molecular weight is decreased and a loss of tensile
strength of the elastomeric material is observed. When cross-
linking occurs, the elastomeric material becomes brittle with a loss •
of elasticity. If no tensil stress is applied, these elastomers can •
be exposed to high concentrations of ozone for long periods of time
without the formation of cracks. However, when stressed as little as •
2 - 3% in extension and exposed to 20 yg/m-* of ozone in extension, |
surface cracks are observed at right angles to the direction of the
stress (Crabtree and Malm 1956). •
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6-17
The major use of these synthetic elastomers is in the production of
tires. Antiozonants are added to the tire formulation and provide a
protective film against ozone degradation (Fisher 1957; Mueller and
Stickney 1970; USEPA 1978). These additives are expensive and their
cost is passed on to the consumer. Other synthetic elastomers are
commercially available which have no olefinic structures and are
chemically resistant to ozone (e.g., silicones, chlorosulphated
polyethylenes and polyurethanes). Although these elastomers are
expensive, they have captured the special-application market
especially for use in hazardous chemical environments.
Dose-response relationships for exposure of elastomeric materials to
ozone have been developed. Unfortunately, most of the work has
involved high ozone levels and elastomeric materials without
antiozonants. Hence, the results do not have an application for
tires in a normal urban environment. An exponential function was
obtained when two styrene-butadiene formulations with several levels
of antiozonant were exposed to 490 y g/m^ of ozone (Edwards and
Storey 1959), The function relates the dose of ozone needed to
produce visible cracks at certain levels of antiozonant.
6.3.4 Effect of Ammonia Pollutant/Material Interactions
There is little information on the effects of atmospheric ammonia on
the corrosion of materials. However, it has been suggested that
ammonia may be a major indirect contributor to the early stages of
atmospheric corrosion (Ross and Callaghan 1966).
Ammonia also plays a prominent role in the atmospheric chemistry of
S02 resulting in the formation of ammonium sulphate aerosols (Bos
1980; Georgie 1970). While there is no information on the effect of
this aerosol on the corrosion of materials, the chemistry of
metal-ammonia complexes would suggest the possibility of such
effects. The primary process would be the modification of stable
corrosion films by the selective dissolution or retention of specific
alloy constituents.
6.3.5 Effect of Particulate Pollutant/Material Interactions
While particulates obviously play a major role in soiling of
surfaces, there appears to be no conclusive correlation between
particulates and materials degradation (Del Monte et al. 1981;
Fassina et al. 1976; USEPA 1981a; Vittori and Fuzzi 1975). In some
cases the particulates serve to increase the effect of other
pollutants by serving as catalysts (Hegg and Hobbs 1978).
However, since particulates probably deliver to surfaces of the order
of 20% of the sulphate and, under certain conditions, up to 50% of
the nitrate (in cities) the role they play in corrosion processes
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6-18
6.4 IMPLICATIONS OF TRENDS AND EPISODICITY
6.5 DISTRIBUTION OF MATERIALS AT RISK
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must not be underestimated (see for example Lindberg and Hosper
1982). Once the dry particles are deposited, wetting events will •
produce the effects described above. I
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Estimates are that atmospheric emissions of NOX will increase in
the order of 35% over the next 20 years (Altshuller and McBean •
1980) while projections show that S02 emissions will remain the •
same or increase only slightly. This means that nitric acid in rain
will substantially increase resulting in increases in potential _
deterioration of materials. •
Episodicity encompasses the variations in levels of air pollutants
occur at a given site. For example, in the northeast, the yearly I
peaks of S02 occur during the winter months. Thus, resulting I
deposition rates may vary through the year.
Episodicity also includes cycles of available moisture, specifically |
cycles of condensation and precipitation. In the northeast, the
specific humidity and temperature of the atmosphere is lower in the
winter. This condition may effect the time of wetness of material I
surfaces in the absence of other sources of moisture. ™
The frequency and intensity of rainfall in relation to dry periods is •
another aspect of episodicity. The erosion caused by the dissolving |
of reaction products (e.g., calcium sulphate) in rain runoff may be
more severe in intense rainfall (Trudgill 1976). Thus, for two M
locations having the same total annual rainfall, the erosion in the I
one with heavier, but less frequent rain events may be greater than
the other, where rain occurs more frequently.
Furthermore, the situation in the episodes between rain events should •
also be considered. In those cases where the damage occurs primarily
by particulate induced attack (e.g., pitting of aluminum), more fl
frequent rain events may wash off the particulates actually reducing |
the overall rate of damage. Variations in the pH of the rain can
influence the solubility of corrosion and weathering products _
(further details available in Section 6.3.2). •
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Most buildings, structures, and statuary subjected to the deterior-
ating process associated with atmospheric transport and deposition •
are found in urban areas. Hence, the spatial distribution will tend |
to follow industrial and demographic patterns, rather than the
sensitive regions identified in aquatic and terrestrial impact _
sections (see Sections 3.5 and 4.5). However, this is not to imply •
that materials at risk are distributed as a simple function of ™
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6-19
population. Attempts to estimate quantities of exposed materials as
a function of aggregate population may lead to erroneous results
(Haynie 1980c; Koontz et al. 1981; Stankunas et al. 1981). Results
from field surveys in two cities indicate that the period of urban
development and the local availability of materials are important
factors determining the distribution of material quantities for rural
areas. Estimates of material distribution have not yet been
attempted.
Inventories of historic structures and monumental statuary have been
undertaken by both the American and the Canadian governments. The
Canadian inventories are maintained by federal, provincial and
municipal agencies and list thousands of structures of cultural
significance. The National Register of Historic Places (U.S.)
includes approximately 25,000 properties. These inventories warrant
closer attention in terms of geographic distribution.
Determination of materials at risk must take into consideration both
the susceptibility of the material as well as the availability and
expense of measures to respond to the incremental damage caused by
air pollution. The principal effect of air pollution damage is to
reduce the length of time the material can serve its intended
purpose. The intended purpose of house paint, for example, is
primarily to protect the underlying wall material. The paint's
expected life may be several years. The purpose of statuary marble,
however, is to display the artistic inspiration of the sculptor and
is intended to last for many generations. Therefore, even though the
paint may be more sensitive to a given level of air pollution than
the marble, the marble statue is more at risk.
Considering the expense and availability of remedial measures,
sculptured stone and bronze are perhaps the most sensitive materials
at risk. Dimension stone is next on the list since it is expensive
to replace once it has been built into a structure. Sheet metals,
brick and block, and concrete lie on the scale somewhere between
dimension stone and surface coatings. The labour and raw materials
involved are less expensive than dimension stone, but more costly
than paints and surface coatings.
6.6 DATA NEEDS AND RESEARCH REQUIREMENTS
Although the examination of the deterioration of materials is a well
established discipline, dose-response relationships, taking account
of atmospheric variables as well as concentrations, are rather poorly
documented. Selection of materials to be investigated should
consider gaps in existing knowledge and significance of materials.
Dose-response relationships need to be better delineated for the
range of pollutants and materials determined in field studies,
controlled environments (including accelerated studies), and in
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laboratory models. There needs to be a study of the roles and
possible damage caused by SO^", SC>2> N03+, N0£ , 03 and particulates, •
both individually and synergistically. Other agents in material I
deterioration need to be studied to determine which constituents are
active on selected materials (e.g., the impact of ammonia on the
corrosion of carbon steels, low-alloy steels and copper alloys), the I
effect of background levels of chloride ions and carbonyl sulphide •
(Gradel et al. 1981) on the corrosion of materials, and the effects
of biological activity on building materials. Further, the role of •
microclimate, including rainfall, cycles and episodes of temperature J|
and moisture, and wind regime needs to be evaluated. Finally, study
should be made to assess mechanical degradation of masonry, soil •
sensitivity and underground corrosion, and stress corrosion cracking •
of Al and Cu alloys.
Many of these needs were discussed in a report to the Electric Power •
Research Institute (Yocom and Grappone 1976). In addition to these •
areas of study there are certain data sets which would help develop
an impact assessment. Pollutant loads should be estimated for •
susceptible structures, giving relative contributions by local |
sources and distant sources. Also, there needs to be more study of
the dose-response relationships for pollutants and materials of _
interest as well as an inventory of the distribution of the •
construction materials and cultural resources sites at risk. Unit ™
cost data for cleaning, maintenance repair and replacement as well as
the human response functions for maintenance and replacement need to I
be developed. •
6.7 METHODOLOGIES
Testing of materials to determine their resistance to atmospheric
corrosion or degradation has been conducted for many years at a •
number of established sites around the world (Committee G-l 1968, for B
metals). Approximately 15 of these sites are in the United States
located east of the 100° meridian with additional sites being •
maintained on a proprietary basis by individual organizations. |
Meteorological and air quality monitoring have not generally been
performed at these sites. Instead, the sites are typically M
characterized as rural, urban, industrial, or marine, to reflect the I
perceived quality of the environment at each site. Recently,
measurements of temperature, rainfall, humidity, wind speed and
direction, solar radiation, S02 and cloride ion concentration, were •
begun at the marine site at Kure Beach, North Carolina (F.L. La Que ™
Corrosion Laboratory) . There has been no report of acidic
deposition, nitrogen oxides, oxidants, particulate matter and ammonia •
being measured at any of the other material test sites. |
Over the years, certain aspects of materials testing in the u
atmosphere have been incorporated into standards by the American •
Society for Testing and Materials (ASTM) to estimate or minimize some
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6-21
of the more obvious uncertainties. The tests for atmospheric
corrosion of metals range from the preparation, cleaning and
evaluation of specimens to the way tests are conducted and data are
recorded (ASTM 1980a). Several methods have been developed to
characterize pollutant levels in the atmosphere. For example,
sulphur dioxide may be determined by using lead peroxide candles
(ASTM 1977) or lead peroxide plates (APHA 1977). A standard for
measuring time-of-wetness for surfaces exposed to the atmosphere has
been prepared in draft form (ASTM 1980b; Sereda et al. 1980). An
ASTM task group was recently established for calibrating the
corrosiveness of the atmosphere at test sites (Baker 1980; Baker and
Lee 1980).
The characterization of time-dependent meteorological air quality
and acidic deposition variables at test sites, and the correlation of
these variables with the response of materials to their environment,
while clearly relevant to atmospheric corrosion and degradation, has
long been recognized as a complex and challenging task. Such an
effort has usually been considered unnecessary where, as in most
cases, the primary goal of materials testing has been to determine
the relative performance of a series of materials and, thereby, to
establish criteria for their selection, improvement, and preservation
in a particular environment. Many studies of this type have been
made on a variety of metallic and nonmetallic materials.
Among the earliest departures from the strategy of comparative
testing were studies led by Larrabee and Coburn (1962), and pursued
on a broader scale by ASTM Committee G-l, to measure the corrosive-
ness of the atmosphere at different test sites for selected metal
alloys. Underlying this interest was the desire for a fundamental
understanding of the interactions between materials and atmospheric
constituents so that the performance of materials could be predicted
based on properties of the material and of the atmosphere. Such a
concept implies a dose-response function which defines the
relationship between the rate of corrosion or degradation and:
(1) the concentration of reactants in the atmosphere and on the
material surface; (2) the nature and disposition of reaction
products; and (3) meteorological and environmental factors which
affect the intensity of exposure to the reactants and the fate of the
products.
The dose-response function quantifies the material-environment
interaction, and provides the fundamental basis for the development
of economic damage functions used for damage (benefit) prediction,
and for designing pollution control strategies (Benarie 1980;
Gillette 1975; Hershaft 1976; Liu and Yu 1976; Mansfeld 1980).
In some laboratory studies of the mechanisms, kinetics, and
thermodynamics of materials corrosion and degradation processes, and
of the effect of specific atmospheric constituents on these
processes, experimental conditions have been well controlled and a
wide variety of sampling and analytical techniques are available
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(Duncan and Spedding 1973a,b; Haynie et al. 1976, 1978; Spence and
Haynie 1974). A typical experimental approach is to vary the dose _
rate of one pollutant while holding other variables constant, and •
study the response of the material. In the case of metals, studies '
of this type are usually done for relatively short times compared to
the time required to form a steady-state corrosion film (ASTM •
Standard Practice, ASTM 1980c for short-term accelerated tests |
methodology). Hence, they are often limited to simple conditions
involving the initiation of corrosion on a bare or slightly oxidized m
metal surface. This approach is adequate for establishing specific •
details of the responses of materials to pollutant dose rates.
However, it has not been effective for describing the performance of
materials in atmospheric exposures, where the permutation and fl
interaction of environmental and meteorological variables is complex •
and constantly shifting over time.
New building components and systems are constantly being designed |
and manufactured. The operating and stress conditions to which they
will be subjected are difficult to predict. Although many tests •
have been developed to accelerate degradation processes of building I
materials, they are seldom fully adequate for reliability predicting
long-term performance. A recommended practice, ASTM (1980c) provides
a framework for the development of improved durability tests. •
Probabilistic concepts have not been applied extensively to materials I
durability problems in the construction industry but these concepts
offer new opportunities for obtaining improved quantitative •
predictions of the service life of building materials in polluted •
environments.
By far the greatest amount of work on atmospheric corrosion and the •
degradation of materials has involved field exposures at regional *
test sites. Here the effects of exposure are clearly defined by
changes in the character and properties of the material (Haynie and I
Upham 1970, 1971; Kucera 1976; Mansfeld 1980; Spence and Haynie 1972; Q
Upham and Salvin 1975). Short- and long-term effects can be
observed; effects in different environments are readily obtained for •
analysis and interpretation. On the other hand, the local •
environmental conditions are obviously variable and it is difficult
to determine cause-and-effeet relationships from regional ^
meteorological/environmental data (Ashton and Sereda 1981; Haynie •
1980a). •
Moreover, the evidence from studies in Europe and North America is •
that the meteorological and materials data obtained at specific sites |
is not generally transferable and applicable to sites at other
locations. Among other reasons, this is because of differences not M
only in the composition of pollutants but, perhaps more importantly, •
a consequence of differences in the properties of the "same
materials". For example, bricks may be made from clay with widely
different chemical composition, may be fired at different tempera- •
tures and for different lengths of time, or may be treated with •
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6-23
different glazes. Finally, when making comparisons between exposures
at different sites, it is important to consider, for example, the
orientation and pitch of the samples and their elevation above ground
(ASTM 1980a). All these variations have an influence on the
ultimate impact of atmospheric components.
Although it is rarely done, it is highly desirable to collect and
analyze the runoff from material exposure samples whenever possible.
These samples provide a direct measure of the amount of material
actually eroded and serve as a check against other techniques such as
measuring the weight loss of the exposure sample. In some cases, the
runoff sample can provide a reliable measurement of material loss
over a much shorter exposure time than possible with other
techniques.
The analysis of field test data to determine the sensitivity of
materials to environmental factors is largely empirical; the funda-
mental reactions and interactions have so far proved to be too
complex to be treated otherwise. Three basic approaches have been
taken for corrosion data. Haynie and others (Haynie et al. 1978;
Haynie and Upham 1971; LeGault and Pearson 1978; Mansfeld 1980; Yocom
and Grappone 1976), utilize a power function, which describes how the
corrosion rate varies over time as the corrosion film ages. The rate
constant is modified by exponential factors, which define the effect
of specific atmospheric constituents. Cramer et al. (1980) employ a
similar approach but use an algebraic factor related to the
composition of the corrosion film to modify the rate constant. In
the second approach, Guttman and Sereda (1968) and others have
expanded the material response function as a Taylor series for a
specific exposure time and determined the coefficients for the
lower-order terms by a least-squares fit of the data. The data do
not generally warrant more than a few linear and interaction terms.
In a third approach, Knotkova-Cermakova et al. (1978) apply feedback
principles to the mathematical analysis whereby the corrosion rate
for the present and all previous times is thought to influence the
corrosion rate in the future by its effect on the growth and aging of
the corrosion film.
Of these approaches, the third appears the most satisfying from a
mechanistic viewpoint. Applications of the first have been quite
useful for extrapolating experimental results. For a given exposure
time, the second approach more readily identifies the important
variables and interactions at a specific site. However, the response
function is nonlinear. Therefore, the results obtained by the second
approach for different test sites are not generally comparable and
should not be used for interpolating to other conditions.
An essential difficulty particularly for heavily used test sites
(e.g., Kearny, NJ.; State College, PA.; and Kure Beach, NC.), has
been the absence of meteorological, air quality and acidic deposition
data which could be correlated with atmospheric corrosion and
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degradation data. Most often, such data have been obtained from a
nearby monitoring or weather station where conditions may not always •
correspond to those at the test site (Haynie 1980a; Haynie and Upham •
1970, 1971; Mansfeld 1980). m
It is now generally recognized that meteorological, acidic deposition I
and air quality instrumentation should be incorporated into field •
materials experiments. In this way, the key atmospheric and
meteorological effects on materials can be determined to provide an •
accurate assessment of the impact of acidic deposition on materials •
corrosion and degradation.
In a study in St. Louis, Missouri, concurrent environmental and •
meteorological measurements were made (Mansfeld 1980). Materials ™
exposure sites were located at nine stations where air quality,
including total suspended particulates (i.e., sulphate and nitrate), I
and meteorological data were recorded. Precipitation chemistry was •
not obtained. These measurements will allow the correlation of
material damage as a function of the recorded environmental •
parameters. I
In a second study begun in 1980, a temporary monitoring station was
established at the Bowling Green U.S. Customs House in New York City. •
This was a joint NPS-EPA contribution to the NATO-Committee on •
Challenges to a Modern Society monitoring project (Livingston 1981).
The objectives of this study were to intercompare site specific •
measurements with the permanent Manhattan monitoring station |
measurements, correlate these measurements with material
deterioration, and investigate a variety of methods to measure stone «
damage. I
A study for investigating material degradation rates over long
exposure periods is being supported by the U.S. Environmental •
Protection Agency. The approach is to measure the erosion of marble 9
gravestones in national cemetaries across the United States (Baer and
Herman 1980). These standard stones, provided by the Veterans •
Administration, are obtained from only three marble quarries, |
providing three sets of chemically uniform indicators. The national
cemetary system provides over 100 exposure sites throughout the ^
country. Since 1873, the stones have recorded the cumulative •
environmental effects at each site.
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6.8 ASSESSMENT OF ECONOMIC DAMAGE
Estimates of the financial losses attributable to air pollution, if •
accompanied by appropriate statements of uncertainty and of assump- I
tions, are useful even if the range of error is fairly large. This
is especially true now in view of increased interest in balancing _
costs of regulation against benefits. Damage cost is a measure of •
material, energy and labour consumption. Premature consumption of *
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6-25
products wastes limited natural resources and consumes labour in
nonproductive tasks. While this may create jobs, it does not contri-
bute to an improved standard of living. Instead, it contributes to
inflation, and hastens the time when certain resources become
scarce.
The task of estimating the damage costs for the effects of long-
range transported air pollution involves many variables. Some of
these variables are difficult or expensive to quantify and some
relate indirectly to damage costs. The values of many of these
variables can only be estimated, because no hard data are available.
In cases where a material is part of a work of art, or has other
cultural value, difficulties arise because of the lack of methodology
for the assignment of an economic value. Some attempts have been
made to estimate damage costs, using conservation and restoration
costs as a surrogate for cultural value. Although this approach
neglects the loss of unquantifiable artistic value originally present
in the structure or object, at least it allows the costs of restor-
ation to be compared with other control measures.
The literature on effects of pollutants on materials describes
various approaches to determining unit costs of extra maintenance,
such as more frequent painting, and earlier replacement resulting
from air pollutants and acidic deposition (Haynie 1980c; Liu and Yu
1976; Yocom and Grappone 1976; Yocom and Upham 1977). These studies
typically involve broad assumptions about the kinds of materials
which are exposed in a given area and are generally based on a
limited variety of materials. No study has produced completely
satisfactory results, and estimates of costs vary widely.
The assessment of economic damage attributable to air pollution
depends on many factors. The rates of deterioration (physical
damage) which can be expected for a material when it is exposed to an
environment which contains known levels of air pollutants and
particulates must be known. The deterioration rate in the absence of
the pollutants must also be known so that the incremental effects of
air pollution (Yocom and Grappone 1976) and the dose-response
relationship can be determined. The distribution of the material in
the environment needs to be catalogued including how the material is
used and whether or not it is protected or exposed. There needs to
be accurate data on pollutant loadings coincident with the material
distribution. Finally, human response to materials damage must be
predicted. In the latter component, there is variability on how and
when to clean, paint, or replace as well as on the selection of
substitute materials which may offer improved performance. There is
also variability as to the extent structures are replaced prior to
the time significant damage would have occurred due to pollutants.
The accuracy of economic estimates is compromised by uncertainties in
all of the above factors.
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The economic assessments available today are crude. In the future
the damage functions for new and old construction materials will be m
determined, and methodologies for determining materials distribution I
will be refined. Then the range of uncertainty in the aggregate cost
of pollution-induced materials damage undoubtedly will be narrowed.
These damage functions and distribution inventories are urgently I
needed. •
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6.9 REFERENCES
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patinas of a gilded horse of St. Mark's Basilica in Venice:
corrosion mechanisms and conservation problems. Studies in
Conservation 21:161-170.
Altshuller, A.P., and McBean, G.A. 1980. The second report of the
United States-Canada Research Consultation Group on the
Long-Range Transport of Air Pollutants. Environment Canada,
Ottawa, Ont.
American Public Health Association (APHA). 1977. Tentative method
of analysis of the sulfation rate of the atmosphere (lead dioxide
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American Society for Testing and Materials (ASTM). 1977. Annual
book of standards. Part 26, ASTM D2010-65: 536-539.
1980a. Annual book of standards. Part 10, ANSI/ASTM
Gl-72: 781-786; ASTM G33-72: 888-890; ANSI/ASTM 750-76:
985-992.
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measurement of time-of-wetness on surfaces exposed to wetting
conditions as in atmospheric corrosion testing. G01.04.01.
(draft)
. 1980c. Annual book of standards. ANSI/ASTM E632-78.
Arnold, A. 1976. Behavior of stone soluble salts in stone
deterioration. In Proc. Second Int. Symp. on the Deterioration
of Building Stones, ed. N. Beloyannis, pp. 27-36. Ministry of
Culture and Science, Athens, Greece.
Ashton, H.E., and Sereda, P.J. 1981. Environment, microenvironment
and the durability of building materials. J. Durability of
Building Materials 1:49-66.
Babick, H., and Stotsky, E. 1978. Atmospheric sulfur compounds and
microbes. Environ. Res. 15:526-531.
Baer, N.S., and Berman, S.M. 1980. Acid rain and material damage in
stone, final report. Submitted to National Atmospheric
Deposition Program, North Carolina Agriculture Research Service,
North Carolina State University, Raleigh, NC.
Bailey, P.S. 1958. The reactions of ozone with organic compounds.
Chem. Rev. 58:925-1010.
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corrosion test sites. Presented at Symp. on Atmospheric
Corrosion. ASTM, Denver CO., 1980. (to be published by ASTM)
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European Community, Environment and Consumer Protection Service,
Paris, France. 65 pp.
and A. Klemin, pp. 140-170.New York:Relnhold Publishing
Corp.
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Baker, E.A. 1980. ASTM Committee G 01-04, Task Group on Calibration
of Atmospheric Test Sites, Minutes of Meeting, May 20, 1980,
Denver, CO. •
Baker, E.A., and Lee, T.S. 1980. Calibration of atmospheric
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Benarie, M. 1980. Environment and quality of life - critical review
of the available physico-chemical material damage functions of •
air pollution. Reoort No. EUR 6643 EN. Commission of the ™
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Bos, R. 1980. Automatic measurement of atmospheric ammonia.
J. Air Pollut. Control Assoc. 30:1222-1224. _
Boyd, W.K., and Fink, F.W. 1974. Corrosion of metals in the *
atmosphere. Report MCIC-74-23, Metals and Ceramics Information
Center, Battelle Columbus Laboratories, Columbus, OH. •
Braun, R.C., and Wilson, M.J.G. 1970. The removal of atmospheric
sulfur by building stones. Environment 4:371-378. m
Building Research Station (BRS). 1970. The assessment of wind
loads. BRE Digest 119, Department of the Environment, U.K. _
Byrne, S.C., and Miller, A.C. 1980. The effect of atmospheric •
pollutant gasses on the formation of corrosive condensate on
aluminum. Presented at Symp. on Atmospheric Corrosion. ASTM,
Denver, CO., 1980. (to be published by ASTM)
Campbell, G.G.; Shurr, G.G.; Slawikowski, D.E.; and Spence, J.W. _
1974. Assessing air pollution damage to coatings. J. Paint •
Technol. 46:59-71. *
Cavender, J.H.; Cox, W.M.; Georgevich, M.; Huey, N.A.; Jutze, G.A.; M
and Zimmer, C. 1971. Interstate surveillance; measurements of P
air pollution using static monitor. Report No. APTD 777, U.S.
Environmental Protection Agency, Research Triangle Park, NC. fi
Copson, H.R. 1945. A theory of the mechanism of rusting of low
alloy steels in the atmosphere. Proc. Am. Soc. Test. Mater. _
45:554-580. •
Crabtree, J. , and Malm, F.S. 1956. Deterioration of rubber from use
and with age. In Engineering uses of rubber, eds. A.T. McPherson •
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Cramer, S.D.; Carter, J.P.; and Covino, Jr., B.S. 1980. Atmospheric
corrosion resistance of steels prepared from the magnetic
fraction of urban refuse. Report of Investigations No. 8447,
U.S. Department of the Interior, Bureau of Mines, Washington, DC.
32 pp.
Del Monte, M.; Sabbioni, C.; and Vittori, 0. 1981. Airborne carbon
particles and marble deterioration. Atmos. Environ. 15:645-652.
Duncan, J.R., and Spedding, D.J. 1973a. Initial reactions of sulfur
dioxide after adsorption onto metals. Corros. Sci. 13:881-889.
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sulfur dioxide onto metal surfaces. Corros. Sci. 13:993-1001.
Edwards, D.C., and Storey, E.B. 1959. A quantitative ozone test for
small specimens. Chem. Can. 11:34-38.
Eurin, P. 1981. Degradation processes of building materials and
components: a short review of some proposals for research. In
Proc. Second Int. Conf. on the Durability of Building Materials
and Components, pp. 354-366. National Bureau of Standards,
Gaithersburg, MD.
Fassina, V. 1978. A survey on air pollution and deterioration of
stonework in Venice. Atmos. Environ. 12:2205-2211.
Fassina, V.; Lazzarini, L.; and Biscontin, G. 1976. Effects of
atmospheric pollutants on the composition of black crusts
deposited on Venetian marbles and stones. In Proc. Second Int.
Symp. on the Deterioration of Building Stones, ed. N. Beloyannis,
pp. 201-211. Ministry of Culture and Science, Athens, Greece.
Fisher, F.L. 1957. Antioxidation and antiozonation. In Chemistry
of natural and synthetic rubbers, ed. F.L. Fisher, pp. 49-55.
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Gauri, K.L. 1979. Effect of acid rain on structures. Presented at
the Symp. on Acid Rain, Preprint #3598, pp. 55-77. American
Society of Civil Engineers, Boston, MA.
Gauri, K.L., and Holden, 1981. Environ. Sci. Technol. 15(4):386-390.
Georgie, H.W. 1970. Contribution to the atmospheric sulfur budget.
J. Geophys. Res. 75:2365-2371.
Gerhard, J., and Haynie, F.H. 1974. Air pollution effects on
catastrophic failure of metals. EPA-650/3-74-009, U.S.
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Guttman, H., and Sereda, P.J. 1968. Measurement of atmospheric
factors affecting the corrosion of met
the atmosphere. ASTM STP-435:326-359.
Hardie, L.A. 1967. The gypsum-anhydrite equilibrium at one
atmosphere pressure. American Minerologist 52:171-200.
Harker, A.B.; Mansfeld, F.B.; Strauss, D.R.; and Landis, D.D. 1980.
Mechanism of S02 and H2S04 aerosol zinc corrosion.
Haynie, F.H. 1980a. Evaluation of the effects of microclimate
differences on corrosion. Presented at Symp. on Atmospheric
Corrosion. ASTM, Denver, CO., 1980. (to be published by
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Gettens, R.J. 1964. Corrosion products of metal antiquities. A
volume from Smithsonian Institute Annual Report - 1963, pp.
547-568. Washington, DC. •
Gillette, D.G. 1975. Sulfur dioxide and material damage. J. Air
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Gradel, I.E.; Kammlott, G.W.; and Franey, J.B. 1981. Carbonyl ™
sulfide: potential agent of atmospheric sulfur corrosion.
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Guttman, H. 1968. Effects of atmospheric factors on the corrosion
of rolled zinc. In Metal corrosion in the atmosphere. ASTM •
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factors affecting the corrosion of metals. Metal corrosion in •
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EPA-600/ 3-80-018, U.S. Environmental Protection Agency, Research
Triangle Park, NC.
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materials confirmed by zinc corrosion data. In Durability of •
building materials and components. ASTM STP-691:157-175. *
1
corrosion damage. Presented at the Int. Symp. on Atmospheric
Corrosion, Abstract 163. The Electrochemical Society, Hollywood,
FL., 1980. ft
Haynie, F.H., and Upham, J.B. 1970. Effects of sulfur dioxide on
the corrosion of zinc. Materials Protection and Performance _
9(8):35-40. •
1971. Effects of atmospheric pollutants on the
corrosion behavior of steels. Materials Protection and •
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1974. Correlation between corrosion behavior of steel
and atmospheric pollution data. In Corrosion in natural
environments. ASTM STP-558:33-51.
Haynie, F.H.; Spence, J.W.; and Upham, J.B. 1976. Effects of
gaseous pollutants on materials - a chamber study.
EPA-600/3-76-015, U.S. Environmental Protection Agency, Research
Triangle Park, NC.
. 1978. Effects of air pollutants on weathering steel and
galvanized steel: a chamber study. In Atmospheric factors
affecting the corrosion of engineering metals. ASTM
STP-646:30-47.
Hegg, D.A., and Hobbs, P.V. 1978. Oxidation of sulfur dioxide in
aqueous systems with particular reference to the atmosphere.
Atmos. Environ. 12:241-253.
Hershaft, A. 1976. Air pollution damage functions. Environ. Sci.
Technol. 10:992-995.
Hicks, B. 1981. Wet and dry surface deposition of air pollutants
and their modeling. Presented at Conf. on the Conservation of
Historic Stone Buildings and Monuments. National Academy of
Science, Washington, DC., 1981.
Holbrow, G.L. 1962. Atmospheric pollution: its measurement and
some effects on paint. J. Oil Color Chem. Assoc. 45:701-718.
ICOMOS. 1967. Conference on the problems of moisture in historic
monuments. Rome:ICROM 1967.
Jakucs, L. 1977. Morphogenetics of karst regions. New York: John
Wiley and Sons.
Judeikis, H.S. 1979. Use of building surfaces in the passive
abatement of gaseous air pollutants. J. of Arch. Res.
7(l):28-33.
Judeikis, H.S., and Stewart, T.B. 1976. Laboratory measurement of
S02 deposition velocities on selected building materials and
soils. Atmos. Environ. 10:769-776.
Judeikis, H.S., and Wrenn, A.G. 1978. Laboratory measurements of NO
and N02 on soil and cement surfaces. Atmos. Environ.
12:2315-2319.
Julien, A.A. 1884. The durability of building stones in New York
City and vicinity. U.S. Tenth Census, 1880. 10:364-384.
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Kotake, S. , and Sano, T. 1981. Simulation model of air pollution in
complex terrains including streets and buildings. Atmos .
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SECTION 7
THE FEASIBILITY OF ESTIMATING THE ECONOMIC BENEFITS
OF CONTROLLING THETRANSBOUNDARY MOVEMENT OF AIR
POLLUTANTS
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7-1
SECTION 7
THE FEASIBILITY OF ESTIMATING THE ECONOMIC BENEFITS OF CONTROLLING
THE TRANSBOUNDARY MOVEMENT OF AIR POLLUTANTS
7.1 INTRODUCTION
7.1.1 Purpose
This section presents a review of the relevant methods of estimating
the monetary value of benefits associated with environmental
protection efforts to deal with the long range transport of air
pollution (LRTAP) problems. An important part of the development of
environmental policy strategies is an understanding of the
relationship of benefits and costs. This section focuses on
techniques to estimate benefits.
The objectives of this section are threefold. The first is to
present an overview of economic methodologies which may be used to
estimate the monetary benefits associated with LRTAP control. In
undertaking this review, the underlying theory is presented and the
applicability and limitations of each technique to the LRTAP receptor
categories are evaluated.
The second objective is to recommend the most appropriate techniques
for deriving monetary values for the increased goods and services due
to reductions in LRTAP. In so doing, the section reviews data
requirements, practicality of the methods, and the extent to which
the methods capture the full measure of benefits.
Finally, the third objective is to present this material in a brief
and readable form, which is readily understood by colleagues in other
disciplines involved in the study of LRTAP effects, but without a
rigorous or extensive knowledge of economics. It is hoped that this
review will provide an understanding of techniques of economic
analysis, and indicate the nature of the information and data needed
from scientific research to apply the economic methods leading to
strategy appraisal.
There is a relatively large body of literature on environmental
economics and benefit/cost analysis. Our purpose here is to
summarize the current state of the art of the valuation of benefits
associated with LRTAP control recognizing that there is sufficient,
if not complete, agreement on this matter. Another Work Group, 3B,
is summarizing state of the art technology and costs.
This review draws upon published works in an attempt to synthesize
theory and application. In particular, Freeman (1979) and Crocker
et al. (1981) have served as important references.
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7.1.2 Background
Neither physical science studies of the effects of deposition alone, •
nor engineering and cost studies of abatement and mitigation
technologies, will suggest appropriate levels for precursor emission _
controls or acidic deposition. Governments and the public are faced •
with choosing among varying levels of damages, effects and costs of ™
control and mitigation, rather than "damages or no damages".
In a world of pervasive markets, prices alone would be sufficient |
means of conveying information about the most appropriate mix of
damages, effects and costs of pollution control. Prices would ^
indicate relative scarcities and provide incentives for allocation of •
resources to the place and time in which they will have the most
value. The most valuable allocation is called, in economic parlance,
an efficient allocation. •
Since there is not a world of pervasive markets, economists, in
dealing with resource allocation issues, attempt to simulate ones •
with benefit/cost analysis. They attempt to assign monetary values, ^
which the gainers and losers would assign, to some change in resource
allocation. The algebraic sum of these dollars is then used in »
determining the necessary level of intervention. If there is a net •
benefit from intervention, then the new resource allocation is said
to be more efficient. The analysis thus shows the benefits of the
intervention to society, and conversely, the costs if steps are not •
taken. m
Consequently, benefit/cost analysis is useful when decision makers |B
want to duplicate the results of a world which reflects individual •
values and preferences. It is limited, however, in that it usually
gives an incomplete accounting of value, and thus it is best seen as ^
an aid to decision making. At a minimum, it constitutes a systematic •
and practical framework for organizing data and for making evalu- ^
ations and comparisons.
Other methods and criteria have been suggested, for assessing the •
environmental effects of resource development projects, or for
evaluating environmental protection strategies such as arbitrary •
weighting procedures, overlay maps, quality and enjoyment indices. •
These are, for the most part, descriptive, and generally do not
provide a consistent, well-developed theory which links human _
preferences and value systems to physical effects being described. •
Moreover, these noneconomic techniques do not provide a systematic ™
and nonarbitrary means of weighting the various physical and
environmental consequences and effects.
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In order to compare the various types of incommensurable entities,
such as changes in crop yield and fish catches, transformations must •
be made to cast these different entities, where possible, into •
comparable units. In addition, the various physical effects must be
given weights, to indicate their relative value to society. Monetary
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7-3
units admirably perform this dual function of providing comparable
weighting units.
A fundamental focus of this paper is to determine the feasibility of
applying monetary valuations to nonmarketed goods, services and
attributes, such as aesthetics, human health and life, ecological
relationships and recreational enjoyment. The difficulty of
assigning monetary values to these environmental attributes is
recognized, the impossibility is not.
It is important to recognize that the monetary value of goods and
services affected by environmental quality is not simply the
willingness-to-pay on the part of the users of the goods and
services. In fact, economists have identified several not neces-
sarily mutually exclusive dimensions of value: (1) activity value
which is derived from the direct use of goods and services affected
by environmental quality; (2) option value which is derived from the
possibility that people might use these goods and services in the
future; and (3) legacy or bequest value which is derived from the
desire of people to leave to their descendants a given level of
environmental quality. An accurate measure of the value of many
goods and services affected by environmental quality would reflect
all these dimensions, where appropriate. For example, the value of
the unique trout fishery in the Adirondacks includes activity value
from those who now use it or would, if, there were not LRTAP; option
value from those who might use it in the future; and legacy value
from those who want to ensure that their children could enjoy the
area in its pristine state in the future.
At this point, a brief discussion of terminology is appropriate. The
terms "damages", "costs" and "benefits" are frequently used inter-
changeably in reference to LRTAP effects on the environment. It is
the choice of a reference point which more clearly determines their
specific meanings. "Benefits" are the gains from preserving existing
environmental quality and from restoring or improving a degraded
area. Since our reference point is a degraded environment, we will
describe the reduction or mitigation of LRTAP effects as benefits.
"Damages" are the mirror image of benefits if, and only if, the path
of environmental degradation is comparable to environmental
improvement. The reference point in this case is a relatively clean
environment, and thus pollution effects constitute "damages" or
"damage costs". Continued or increased emissions, in a somewhat
polluted environment, are also likely to have effects which would be
considered as damages.
For the purpose of this section, we have attempted to be consistent
in our use of the terms. The word "costs" is used primarily in
reference to LRTAP control or abatement efforts. Benefits are the
gains associated with pollution reduction or prevention, given our
reference point of an environment already affected by pollution.
Our economic measure of benefits is, therefore, the value which
people place on reducing the effects of LRTAP, and our purpose is to
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7-4
In the quadrant IV, a benefits function can be drawn which relates
the dollar value of benefits to particular LRTAP emission levels by
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indicate how monetary values might be assigned to the physical
effects resulting from LRTAP abatement. mm
7.1.3 Emission-Benefit Relationship
The relationship between residual emissions, (e.g., SOX and NOX), w
and monetary benefits is complex, and varies among receptor
categories. However, the general relationship consists of three •
primary linkages: (1) the relationship between emission discharges |
and ambient environmental quality; (2) the relationship between
ambient environmental quality and the direct and indirect effects on ^
people (the dose-response); and (3) the relationship between direct •
and indirect effects on people and the economic value of these
effects (user value). The linkages between the various elements can
be represented in a simple quadrant diagram in Figure 7-1. I
Quadrant I shows a transformation function which relates emission
levels to deposition (used here as inverse for environmental •
quality). In the case of acidic deposition, long range transport •
models are being used to show the spatial and temporal relationship
between S02 emissions and sulphate and total sulphur deposition. _
Quadrant II shows a functional relationship between deposition (or •
environmental quality) and activity levels. This relationship
between emissions and activity levels, is very complex and varies •
considerably with the receptor category. |
In the case of sports fishing, the relationship in the diagram is a mm
gross simplification of the linkage between sulphur deposition and •
days of sports fishing activity. There is actually a complex affect
on the aquatic ecosystem. The amount of change in pH (and metal
ions) varies with the buffering capacity (sensitivity) of the •
waterbody. The change in pH will have various effects on the fish •
population (i.e., rough , warm and cold species). Finally,
recreational fishing may be affected by the changes in species types •
available as well as a reduction in the number of days fishing is I
permitted. This assumes that the stock of fish is independent of
fishing pressures. _
Where the effects are direct (e.g., human health) the function in ™
quadrant II illustrates the relationship between ambient
environmental quality (e.g., sulphate concentrations) and human •
mortality. |
Quadrant III relates activity to dollar values. The relationship is mm
direct; as activity increases, total economic value increases. This •
function could illustrate the relationship between activities and
both its primary and secondary economic value.
(B
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7-5
Activity
Benefits
Costs
IV
Figure 7-1.
Deposition
Emissions
Conceptual relationship between emissions and economic
effects.
Activity
$$$ L.
IV
Deposition
Figure 7-2.
W
Variation in effects due to different emission-
deposition relationships.
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following a given level through each quadrant. Thus, the total
benefits are the integral under the curve. In addition, this
quadrant can be used to show a control cost function (shown in dashed •
lines in Figure 7-1). •
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Actually the functions are not as straightforward as shown in this
quadrant diagram. There are elements of uncertainty, which may
result from inaccuracies in measurement or from the use of
assumptions drawn from insufficient data. •
For example, the best estimate of the relationship between emissions
and deposition (i.e., the transformation function) is represented by
a range. The upper limit is A and the lower limit is B. The •
benefits range in quadrant IV is no longer a curve. It is now the •
area WXYZ (Figure 7-2) where the range of benefits at emission level
"a" is Z to Y. If the functions in quadrants II and III are •
similarly presented as ranges, the total benefits area in quadrant IV |
becomes an even larger area.
In the case of LRTAP, this section reviews the economic methodologies I
for assigning monetary values to benefits for the activity categories ™
shown in Table 7-1.
A critical link in this process is the relationship between dose and •
response (Quadrant II). Until more consistent research data are
forthcoming which relate damages to various levels of LRTAP, it will •
be difficult to provide reliable estimates of the economic values of •
the benefits of LRTAP reduction.
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7.1.4 Efficiency and Equity Considerations
The physical damages of LRTAP result in reduced economic benefits for
society. Conversely, the reduction of these damages offsets these
welfare losses, which may not be shared equally by all members of
society. The decision by government to intervene or not to intervene ^
thus has both efficiency and equity implications. •
It is often possible to reallocate resources, (e.g., spend more or
less on environmental protection) to increase the net value of I
production or output to society. This output encompasses goods and V
services provided by the environment as well as those produced by man
and sold in markets. Increases in this net value of output result in •
an increase in economic welfare. V
Changes in economic welfare associated with environmental damages or —
environmental protection activities sometimes may be measured by •
changes in the monetary values assigned by all individuals affected ™
by an action. Thus, a reallocation of resources and efforts will
increase efficiency if it results in an increase in the social value B
of goods and services produced by the economy (or by natural |
environments), as indicated by individuals' demand for them.
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7-7
TABLE 7-1. ACTIVITY CATEGORIES
1. Aquatic
2. Terrestrial
a. commercial
b. recreational
c. ecosystem
a. agricultural
b. forest
c. ecosystem
3. Man-made buildings, structures and artifacts
4. Water Systems
5. Human Health
6. Visibility
a. materials
b. historic
a. treatment
b. materials
a. morbidity
b. mortality
a. aesthetic
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The equity impact relates to the redistribution of economic welfare
amongst individuals. These individuals can be further divided into
groups, regions, countries, or generations. The simplest way in •
which to view this impact is to examine the changes in the distri- j|
bution of monetary income which result from alternative strategies.
A wide variety of individuals of differing nationality, income and •
social class, and generation, are potentially affected by LRTAP *
effects and LRTAP abatement alternatives. Some individuals may be
more significantly affected than others by social choices to abate or •
not to abate LRTAP. They may be located in source or sensitive •
receptor regions or they may have a preference for goods and services
related to air quality in these regions. A decision to maintain A
existing air quality management practices produces gains to Jj
individuals who use the environment for production (e.g., an employee
or shareholder of firms in source regions) and consumption (e.g., use _
of automobiles in source regions) purposes or consume the products 8
and services of LRTAP-source firms. Damages are imposed on those
individuals who use the environment for production (e.g.,
agriculture), consumption activities (e.g., water-based recreation in •
acid-sensitive lakes) or consume the products and services of LRTAP- •
affected firms. The roles of gainers and losers are reversed if
LRTAP abatement is contemplated. •
The adverse effects of pollutants result in the involuntary surrender
of rights to both common pool resources (e.g., airsheds) and _
individual property, or the usurpation of these rights. For example, •
residents in both Canada and the United States involuntarily give up ™
some of their rights to clean air and lakes with fish populations to
the coal-burning utilities and the nonferrous smelting industry. H
A political jurisdiction may surrender these rights from one area to V
another if it is determined (by concensus or voting) that the
collective good of the nation or region is enhanced. However, A
regions or nations may not share in the collective good that may •
result from the transfer. There is no forum for arriving at a
concensus between two nations other than negotiation and bargaining.
Thus, redistribution effects and property rights among groups, ™
regions, or generations may have significance for social welfare,
depending upon the relative weights attached to individuals, •
countries, or generations, and whether or not compensation is |
actually paid to those affected by the changes.
The compensation principle adds further complications to the property •
rights issue, because it is unlikely that compensation will be paid
between losers and gainers. A benefit/cost analysis shows a
particular course of action worthwhile if the gains would be I
sufficient to compensate the losers. If compensation does not take •
place and the distributional weights are important, efficiency
conditions may not be satisfied. Thus, the lack of mechanisms or •
incentives to preserve or compensate the rights of others may result •
in an economically inefficient allocation of resources.
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7-9
The assignment of implicit or explicit welfare weights has been
subject to significant criticism. It has been noted that data and
analytical problems severely limit the ability to trace the eventual
distribution of economic effects to individuals, as tax payers,
resource suppliers, and consumers. The development of implicit
welfare weights from past governmental decisions also requires the
assumption that elected officials had full knowledge of the magnitude
and composition of the economic effects, when decisions were made
(Freeman 1969). The ability of elected government officials to
generate an optimal set of equity weights which would be stable over
time has also been questioned (Steiner 1977).
There is disagreement among applied welfare theorists, as to the
treatment of distributional effects. A display of the distributional
consequences of social choices is recommended as a supplement to the
statement of economic efficiency impacts (Haveman and Weisbrod 1975;
McKean 1958; Mishan 1971). Some analysts have also recommended that
the display process be taken a step further. A series of welfare-
weighted calculations could be used as an alternative to the
weighting functions (Eckstein 1961). The distributive consequences
of alternative weights can then be easily identified. Freeman (1969)
outlines a more formal process whereby each government agency spells
out program objectives and recommends weighting functions. These
weighting functions are then reviewed and approved by the central
budget agency (e.g., Office of Management and Budget, or Treasury
Board) to ensure weighting consistency among programs relative to
overall governmental priorities. Present federal project evaluation
procedures of the Canadian Treasury Board (1976) and U.S. Water
Resources Council guidelines (1980) basically conform to the display
format for distributional consequences recommended in the literature.
Therefore, the redistributive effects are associated with most of the
benefits of LRTAP control and these should also be taken into
consideration.
7.2 BENEFITS: CONCEPTUAL APPROACHES
This chapter distinguishes between primary and secondary monetary
benefits associated with changes in activities. Primary benefit is
the willingness of society to pay for goods and services resulting
from changes in environmental quality, or the compensation required
to restore welfare to original levels. The willingness to pay on the
part of consumers is described graphically as the area under a demand
curve. Consumer surplus is the difference between willingness to pay
and actual expenditures. It is the change in consumer surplus
resulting from changes in LRTAP which provides a useful measure of
benefits to consumers. The willingness to supply on the part of
producers is described graphically as the area under a supply curve.
Producer's surplus is the difference between the price line and the
supply curve and it is the change in producer surplus due to LRTAP
which provides a useful measure of benefits to producers. (For those
unfamiliar with economics, see the Appendix which gives a few key
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7.2.1.1 Market Approach
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concepts of economic theory). Most of this section will describe the
three basic approaches (i.e., market, imputed market and nonmarket)
for estimating willingness to pay on the part of consumers and •
producers. Secondary monetary benefits are changes in the levels of "i
economic activity among regions. Some will gain while others lose.
These are not usually contributions to national economic efficiency, •
but are transfers from one region to another. Nonetheless, they are j|
important at the regional level.
7.2.1 Primary Benefits ™
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In many cases, the value of a change induced by an improvement of
environmental quality can be ascertained by direct observation of a •
change in the market. For example, a reduction in LRTAP which •
results in an improvement in environmental quality may produce
increases in crop yields and commercial fish catches. The benefits —
of a change in environmental quality include the total value of price •
changes to consumers and net income to producers. In the following •
paragraphs, we will present a few techniques for estimating these
changes. •
The Net Factor Income approach provides a measure of the change in
producer's income resulting from change in output due to improved _
environmental quality. One way to measure changes in producer's •
income is to specify the dose-response relationship so that the
change in output (as measured by yield, catch, etc.) can be
determined. Assuming that the increase in output does not affect •
price, the change in income is calculated by multiplying output •
change by the relevant market price. If the output change is
sufficiently significant that price falls, the new price should be •
used. |
An innovative application of the Net Factor Income approach is to _
determine the change in input costs rather than output. Changes in •
producer's income are measured by the difference in the cost of ™
production to sustain the same yield at different levels of LRTAP
deposition. The advantage of this approach is that estimates can be B
made in the absence of specific dose-response data. 0
Partial Budgeting, a technique similar to net factor income, im
estimates the extent of benefits from environmental improvement by •
calculating the effects on key portions of a budget of a represen-
tative (e.g., farm) enterprise. Expansion factors can then be used
to extrapolate from the effect on the enterprise to the effect on the •
industry, or on a specific crop or kind of livestock. Examples of •
this technique are in use in Economics and Statistics Services, USDA.
Economic gains are estimated from production and sales of a certain •
crop because of reduced insect damage, or agricultural losses to •
farms and ranches from strip mining of coal.
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Perfect Substitutes is another technique to be used where a change in
environmental quality requires less of other productive inputs. This
situation is best illustrated where decreases in LRTAP deposition
result in reduced use of lime to treat drinking water or maintain
lake water pH. The value of benefits is measured by the cost savings
for lime.
There are some important qualifications, which affect the extent to
which these market methods incorrectly estimate the value of
benefits. The first of these problems concerns the use of "partial
equilibrium" analysis in making estimates of price and quantity
changes. All other variables relevant to demand and supply of goods
and services are assumed not to change. However, the availability of
substitutes is an important determinant of demand, which may well be
altered by the effects of acidic deposition. The a priori
implications of change in these aspects can be noted, but the
empirical verification of these hypotheses in some cases is
difficult. The market price may generally be used as the relevant
measure of unit value. However, where government support or other
policies which raise prices is in effect, use of this (supported or
other) price may overstate the value of benefits.
7.2.1.2 Imputed Market Approach
Where there are no organized markets for the goods or services of the
environment (e.g., visibility) or for goods affected by quality
(e.g., recreation), a number of imputed market approaches are
available for deriving or inferring their monetary value.
The Property Value Method is an imputed market approach which has
been used to value environmental benefits. Economists have long been
interested in the relationship between property values and levels of
environmental quality. It is suggested that variations in the level
of environmental quality will affect the value of otherwise similar
properties (Ridker 1967; Ridker and Henning 1967). The value
obtained should reflect tangible and intangible values of
environmental quality, insofar as these are perceived by individuals
in the property market. This facet may pose problems in the case of
LRTAP. Since these effects are not well understood, the property
market may not accurately reflect the adjustment. Therefore, the
property value method is not considered an appropriate method to
measure the benefits of LRTAP reduction.
Hedonic Price or Demand Analysis is a more general application of
this specific property value technique. A demand function for a
public good is estimated through a two step procedure. First, the
implicit price of environmental quality is estimated based on
property value (in the case of visibility) or travel cost (in the
case of recreation). Then the implicit prices are compared with
variations in environmental quality to determine a demand function.
However, only under a rather broad set of circumstances is the demand
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The major issues involved in these techniques relate to three
categories of bias: hypothetical, strategic, and information. They
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function derived as an accurate measure of willingness to pay for
environmental quality.
Complementary Expenditure is yet a third imputed market technique. •
Variation in travel cost is the most frequently used expenditure and
is the basis for imputing the value an individual places on a •
recreation experience. The maximum expenditure of an individual is |
the basis for deriving a demand curve. Other expenditures (e.g.,
expenses for fishing equipment) may be used to impute values. _
Risk Premium is the imputed market technique most frequently used in
valuing mortality. These risk premiums may include wage
differentials among occupations and insurance premiums. The •
technique could possibly be used to value changes in morbidity as •
well, but foregone wages and medical costs are the more common
valuation approach. •
7.2.1.3 Nonmarket Approach
A third major approach to valuation is a direct enquiry to •
individuals about their willingness to pay for changes in ™
environmental quality. Information is obtained through an interview
method, whereby respondents are asked to reveal their preferences for •
environmental services. In some cases (e.g., visibility changes), ^
this approach is an alternative to another major approach (e.g.,
imputed market). In cases where no markets exist for estimating «
option value, it is the only approach for valuation. •
The Bidding Game is a nonmarket approach to preference revelation.
The technique consists of constructing an artificial (i.e., •
contingent) market and simulating market transactions. The m
interviewer/auctioneer presents the respondents with a set of
possible states or contingencies for the relevant environmental K
service. The respondent is asked to assign a price or is asked •
whether he or she concurs with a price for each possibility. The
auctioneer enters into a bidding process to determine if the _
respondent would pay (receive) a higher (lower) price than that •
stated initially. The process continues until the auctioneer has ™
determined the highest (lowest) bid.
Rank Ordering is a second nonmarket technique requiring similar |
information about hypothetical environmental situations to the
bidding game. Also some measure of the costs or price of a visit was M
required. Individuals are asked to rank the hypothetical situations •
from the least to the most desirable. These rankings reveal
tradeoffs among the environmental services, other attributes of an
area and price of admission. These tradeoffs form the basis for •
estimating the value of various levels of environmental quality. V
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7-13
are summarized in Rowe and Chestnut (1981), but are briefly discussed
here.
Hypothetical bias refers to the fact that the hypothetical nature of
the question may not ellicit accurate bidst In order to minimize the
hypothetical bias, it is essential that the scenario which is defined
for the respondent be as real and credible as is possible.
Otherwise, the respondents may feel that the game, being totally
hypothetical, is not really relevant. In this case, they may not
treat the interview seriously and hence fail to reveal true bids.
Strategic bias is introduced if the respondents have an incentive to
conceal their true preferences in order either to avoid payment (be a
free rider) or to influence the outcome. Strategic behaviour occurs
when respondents attempt to impose their preferences on others by
bidding in such a way as to influence the mean bid and hence the
outcome (Brookshire et al. 1976). The survey must be designed so as
to minimize the influence of these biases.
Information or starting point bias refers to the extent to which the
information provided to respondents may influence their bids.
7.2.2 Secondary Benefits
The approaches described above are techniques for the valuation of
the primary (efficiency) benefits associated with a particular
commodity or service. However, the benefits of environmental
improvement are not limited to a particular good. Changes in income
to producers in terms of crop or forest products could have important
impacts on jobs and income in regions where these activities are a
part of the economic base. Similarly, changes in sports fishing
activity will affect the overall sector. These effects will be more
significant in areas where these activities form a greater proportion
of the economic base. Since these secondary benefits will accrue to
different economic sectors and to various geographic regions,
analysis on a sectoral/regional scale provides an additional measure
of welfare change.
Various regional economic analysis techniques are available to
account for secondary benefits. They measure the effects on spending
and respending patterns (multipliers) and the direct, indirect, and
induced effects on other economic sectors (linkages) and on the
overall level of economic activity in the region (Bender 1975;
Conoposk 1978).
Although the change in the overall level of economic activity due to
LRTAP may be a transfer from one region or country, the effect of the
transfer is important to that region or country. Thus, an analysis
of these secondary effects is an important part of the total estimate
of benefits.
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7.3 BENEFIT ESTIMATION TECHNIQUES
7.3.1 Aquatic
7.3.1.1 Recreational Fishery
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The literature on benefit estimation techniques offers several •
approaches for transforming effects of LRTAP on various receptors •
into dollar values. These have been broadly categorized into three
groups by their reliance on market, imputed market and nonmarket M
mechanisms for the generation of value information. This section •
examines the appropriateness of each technique for the various LRTAP
receptor categories and the methodology employed. It comments on the
technique's specific application to the valuation of benefits •
associated with LRTAP abatement for each of the identified receptor •
categories.
In considering which techniques are most appropriate to LRTAP, jf
several factors are relevant. The benefit estimation techniques
should have a solid theoretical basis (i.e., consistent with economic •
theory). Although the techniques have theoretical consistency, it is •
important that the empirical implementation be practical in terms of
data and computational requirements.
The techniques selected should minimize the degree of uncertainty and m
contentiousness associated with the results. Application of the most
appropriate technique to a receptor category and a clear statement of
assumptions and conditions are essential for credible results.
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The results must be obtained at a reasonable cost. For example, _
survey design must balance practicality and cost with reliability of •
results. It is not feasible to interview every member of society. ™
Instead, it is necessary to interview representative sample groups
drawn from the population. •
These criteria are not independent of each other. An analyst is
forced to make tradeoffs or concessions. Ultimately, it is desirable »
to obtain the most useful results at reasonable cost for the purpose •
at hand. *
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The conceptually correct procedure for estimating the value to users
of changes in recreational fishing is to estimate the willingness- «
to-pay for each site. Willingness-to-pay is a function of I
socioeconomic characteristics, the quality of the recreational ™
experience at the site and at substitute sites, and the (travel and
other expenditure) cost to get to the site and substitute sites •
(Freeman 1979). V
A demand curve which graphically summarizes willingness-to-pay for an •
individual site would relate the number of fishing days to the prices •
of this experience assuming no changes in income and tastes. The
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7-15
demand curve for a given lake is given in Figure 7-3. If the price
of admission to the lake is OA per day, the quantity consumed will be
OX^. The value of the recreation site given the initial level of
water quality is the consumer surplus as measured by the area ABC.
If no admission is charged (i.e., price is zero), then consumer
surplus is the area under the demand curve, (i.e., OBCD^).
Now, if acidic deposition is reduced and water quality improves, the
demand curve will shift to the right (02). The net economic
benefit is the increase in willingness-to-pay as measured by the area
between the two demand curves, BDEC. The net benefit is the
increased value to existing users as well as the value to the new
users.
The problem with this theoretical approach is that most recreational
fisheries charge a zero price or nominal entrance fee. Where there
is no fee or variation in market price, a market approach to
valuation is impossible.
One imputed market approach is the Clawson-Knetsch travel-cost
method, which infers behavioural responses to price changes reflected
by differences in travel cost. This method is site specific and
estimates demand for a particular park or lake and not the fishing
activity itself. This method is done by circumscribing a
recreational fishing site with a series of concentric zones. For
each zone, travel costs and visitation rates are calculated based on
a survey of the origin of visitors.
Four limitations of this technique are the assumptions that all
travel costs are incurred for the purpose of visiting the specified
site, that travel time can be correctly valued, that only present
users are accounted for, and that response to a change in quality
cannot be inferred from a single site. The second can be handled by
using a shadow price to value time but will result in different
values depending on the assumptions used. The third can be handled
by surveying non-users within the region to determine their response
to water quality improvements at a given site. The fourth requires
data on a number of sites of varying quality. However, this method
is more complex if there are several recreation sites which are
substitutes such as is likely with the effects of LRTAP. In this
case, the demand for any one site is a function of prices and
distances to other competing sites.
A second imputed market approach, the property value technique, has
seldom been applied to changes in water quality due to the
difficulties in correctly accounting for the interaction of water
quality and distance from water on property values. Also it ignores
the value of changes in water quality to recreationists who do not
own property in the vicinity of the lake or stream. This technique
is not recommended for a valuation of benefits of environmental
improvement for a recreational fishery for these reasons. In
addition, data for other "imputed markets approaches" are more
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Figure 7-3. Change in demand due to water quality improvement.
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readily available and it is doubtful that in the absence of solid
information on dose-response relationships the property market would
be able to reflect adequately the pervasive effects of LRTAP.
A third imputed market approach is the participation model. The
technique relates participation in specific recreation activities by
a given population to the socioeconomic characteristics of that
population and to the supply of recreation opportunities available to
develop estimates of quantity. If participation equations for
specific populations are estimated, it is possible to predict the
increase in participation to be expected with an increase in fishing
opportunities or ambient water quality. The value of a recreation
day of particular type must be inferred by other methods, (e.g.,
travel cost). Then one component of recreation benefits (i.e., the
value to new users) can be estimated by multiplying the increase in
recreation days (quantity) by the assumed value per day (price)
(Figure 7-3). This approach is limited in that it does not capture
the utility associated with the current level of use (quantity Xj),
that is the area BDFC in Figure 7-3, and there is considerable
uncertainty about the value assigned to a recreation day.
A probabilistic participation model is perhaps the best technique for
accounting for a broad range of locations, accessibilities, and fish
types affected by changes in reductions in acidic deposition and the
interacting adjustments made by recreational fishermen to the
additions of available water bodies (Russell 1981). This method
requires:
1. Estimation of two types of probability-of- participation
equations:
a. The probability that a randomly chosen member of the U.S.
or Canadian population is an angler at all;
b. The probability that a randomly chosen angler spends at
least some time in a year fishing for a particular species.
2. Estimation of an equation predicting the number of days of
fishing per year engaged in by anglers for various fish
species.
The equations could be estimated using data on existing water availa-
bility. Next, changes in the probabilities and days of participation
could be projected for a condition under which additional bodies of
water were made available for fishing due to reductions in acid
deposition. Fishing days in this future state, less fishing days in
the present case, would give the projected increase in fishing by
type of fishing. The increase in days for each type of fishing would
then be valued at the appropriate figure for consumer surplus per
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Species
Type
Lake Type
Susceptible Very Susceptible
(little buffer) (no buffer)
Fish
Population
Surviving
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day. However, the application of this technique to the benefits of
reduced LRTAP effects is complex given its data requirements. _
The major limitation in applying this technique is in relating
changes in total sulphur or sulphate deposition to changes in water
available by species type. The aquatic effects work group would have I
to provide an inventory of existing water bodies, differentiated by m
susceptibility to acidic deposition and general species type (rough,
warm water and cold water). In addition, they would have to provide •
dose-response data, such as different deposition levels (kg of wet •
sulphate per ha per year) that would permit the survival of fish
populations by species type for each water regime in the following _
format: •
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rough <15 + 5 kg/ha/yr <10 + 4 kg/ha/yr >90% I
— — — — — v
warm water £15 + 5 kg/ha/yr <10 + 4 kg/ha/yr _^90%
cold water <10 + 2 kg/ha/yr £5+2 kg/ha/yr >90% •
Without this type of information, or another similar approximation, |
we cannot accurately estimate the economic value of a change in
recreational fishing due to a change in acidic deposition. •
For recreational fisheries a contingent market approach using
surveys can be used. The approach attempts to elicit values of how
respondents think they would behave if a proposed water quality •
change were to occur in a hypothetical situation. There is some •
skepticism about these approaches primarily because they assume that
individuals are capable of predicting and willing to predict •
accurately their response behaviour to a hypothetical situation. In |
addition, the accuracy of the results may be questioned due to
information, strategic or starting point biases as discussed in _
Section 2. •
There are two limitations to the above techniques. They fail to
capture option and legacy values of the recreational fishery or the •
aquatic ecosystem. They also place emphasis on valuation of a •
particular activity rather than the economic sector (i.e., recreation
and tourism), which is based on recreational fishing (i.e., secondary
benefits).
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7.3.1.2 Commercial Fishery
The conceptually correct approach for valuation in the case of
commercial fisheries requires estimation of producers' responses and
market effects. In the simplest case, assume that only a few
commercial fishermen benefit from a reduction in acidic deposition.
If the fishery is being appropriately managed to maximize net
economic yield, the benefit of reducing acidic deposition is equal to
the market value of the increased yield. This is net of any changes
in expenditure on variable factors of production. For existing
fisheries, there would be little variable cost change outside of
shifting fishing grounds or shifting from less desirable to more
desirable species.
In a more realistic case, the value of restoring a commercial fishery
will depend upon several variables unique to a given situation. If
there is free access to the fishery, producer surplus would accrue to
the existing fishermen only in the short run. This surplus would
attract additional fishermen to the fishery, which would reduce it to
zero. If the change in the commercial catch is significant, then
there would be a need to estimate price effects (consumer surplus) as
well.
In addition, restoration of a commercial fishery in economically
depressed areas could conceivably be sufficient to strengthen a
region's economic base and hence income and employment. Maintaining
the regional population would prevent negative external effects upon
the rest of society. This is because there are higher levels of
congestion in urban areas due to migration from depressed fishing
areas. Some of the secondary benefits can be estimated using a
regional income and employment model.
The specificity of the dose-response relationship between reductions
in LRTAP deposition and increases in commercial fish populations and
catches will determine the reliability of any estimates of the
benefits to this economic sector.
7.3.1.3 Aquatic Ecosystem
Any valuation of the benefits of reducing acidic deposition should
reflect the value of all changes in the aquatic ecosystem, rather
than just recreational and commercial fisheries. Changes in
salamander and loon populations could result from reduced acidic
deposition, and these changes could affect activity, option and
existence values. Limitations in dose-response functions and the
absence of economic studies in this area will make it difficult to
measure these values. Insofar as they are excluded, the total
benefits will be underestimated. These changes should therefore be
stated in qualitative terms.
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7.3.2 Terrestrial
7.3.2.1 Agriculture
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The conceptually correct procedure for valuing changes in the
agricultural sector is the determination of changes in producer's and .
consumer's surplus due to a change in environmental quality. The •
changes in the surpluses depend upon costs of production, demand, and *
market structure. Since knowledge of these parameters suggests that
most of the benefits of reducing LRTAP effects will accrue to I
producers, benefits may be estimated from observed or predicted •
changes in net income of certain factor inputs. The change in net
income accrues as profit to the farmer increases or as surplus income
increases above the fixed factor of production.
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In the case of acidic deposition, there is limited information about
effects on agricultural productivity. A satisfactory estimate of the •
change in net factor income can be obtained from the product of ™
changes in crop yields multiplied by market prices. For this
procedure to be a reasonable estimate of benefits, there must not be
government intervention to support the price of affected crops nor
changes in expenditures for other production inputs. In addition, it
is assumed that producers have neither undertaken mitigation «
measures (e.g., liming), nor changed cropping patterns in response to •
acidic deposition.
If acidic deposition is shown to have a measurable impact on some •
crops across a large geographic area, then it is recommended that •
consideration be given to changes in prices as well as yield. Given
that many agricultural crops have inelastic demand curves (i.e., a •
small change in quantity demanded results in a larger proportionate f
change in price), accounting for price effects would considerably
improve the total estimate of benefits and indicate the distribution _
of benefits between producers and consumers. In the absence of dose- I
response data, an alternative estimate of net income changes due to •
differences in LRTAP is nonetheless possible. This approach requires
measuring the difference in the cost of producing a given level of I
agricultural output under different environmental quality |
conditions.
7.3.2.2 Forestry I
The conceptually correct procedure for estimating the value of
changes in the forestry sector is similar to the agriculture sector •
where the market approach is used. Since knowledge of production, ™
demand, and market structure suggest that the benefits of reducing
acidic deposition will accrue to producers, benefits may be estimated
from observed or predicted changes in net factor income.
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In the case of acidic deposition, a frequently satisfactory technique _
to estimate the change in net factor income is to use the change in •
timber yield multiplied by the market value differentiated by species
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and product. Moreover, it presumes that the demand for forest
products is highly inelastic (i.e., insensitive to price changes).
The procedure is less complex for the forestry sector than in the
agriculture sector, because there is less government intervention.
Also there are fewer known uses of mitigation measures or changes in
tree planting patterns due to acidic deposition.
7.3.2.3 Ecosystem
Any valuation of the benefits of reducing acidic deposition should
reflect the value of all changes in the terrestrial ecosystem, not
just agricultural and forestry activities. Change in nutrient
composition of soils is a major change which may not be immediately
captured by changes in yields in the agriculture and forestry
sectors. This change, as well as changes in terrestrial animal
populations, would have some affect on activity, option and existence
values. Although these are best measured by means of a survey, it is
unlikely that individuals would be able to assign accurate options or
values to the terrestrial ecosystem considering the dearth of
dose-response information.
7.3.3 Water Supply
The conceptually correct procedure for valuing a reduction in the
direct effects of acidic deposition on water supplies, is the
reduction of treatment cost. These changes in treatment costs are a
first approximation as long as they do not change other forms of
producers activities, cause substitutions among factor inputs, or
change prices of outputs.
Although the use of changes in treatment costs is recommended as a
benefit measure, there may be problems in making an empirical
estimate. The problem lies in correctly assigning a percentage of
liming costs to the mitigation of acidic deposition effects. Even if
there were no acidic deposition, industries and municipalities would
probably continue their current treatment practices of balancing the
pH of water. Consequently, we would provide at best only an upper
bound on benefits by assigning all liming costs in areas of high
atmospheric acidic deposition.
7.3.4 Effects on Buildings and Structures
The conceptually correct procedure for valuing the reduction in the
effects of acidic deposition on commonly used materials, is the
annual equivalence of the present value difference in life cycle
costs of production processes. The difference is appropriate for
reductions in deposition which extend the useful life of materials
(including water supply systems), reduce maintenance or repair costs,
or eliminate the need for higher initial costs for damage resistant
materials.
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The recommended approach would use the annual equivalence of
difference in life cycle costs for commonly used material. Annual
equivalence would be calculated in a two step procedure (Maler and •
Wyzga 1976). The first step estimates the present value of the •
difference in the stream of current replacement costs before and
after acidic deposition. The second step calculates the annual
equivalent flow of the present value of the reduction in damages.
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Benefit estimation for commonly used materials requires information »
about changes in rates of deterioration (dose-response) or •
maintenance, and distribution of susceptible materials. Thus,
information is essential to the determination of the value of
benefits. •
While the life cycle cost approach values the benefits of reduced
repair and replacement costs, it does not capture the historical •
value of buildings and monuments. This is an intangible, nonmonetary •
value, which can best be determined by a willingness-to-pay survey of
viewers for the aesthetics of less damaged structures and statues.
However, this method will result in an underestimate of total value •
in that it fails to capture option and legacy values. A second and •
important limitation of the contingent valuation method is due to the
lack of a proven approach. Although surveys have been tested and •
validated in other areas (e.g., recreation and visibility), f
additional research would be required prior to their application to
derive historical values. •
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7.3.5 Human Health
7.3.5.1 Mortality
An understanding of the dose-response relationship between air •
pollutants and mortality and morbidity is needed to value changes. •
Animal and clinical studies provide a basis for confirming a
relationship between air pollution and health. Some epidemiological
studies estimate a dose-response between air pollutants and mortality •
and morbidity. Epidemiological and clinical studies can therefore be "
used to indicate the probability or risk of mortality or morbidity
under different environmental conditions. These types of data must B
also be matched with changes in population exposed to determine 0
changes in mortality or morbidity.
The amount that an individual must be paid to accept additional risk •
is conceptually the correct procedure for estimating the value of
human life. When aggregated over many individuals, this willingness-
to-pay, is usually referred to as the value of statistical life, or •
the value of a statistical death avoided. It is simply a shorthand •
way to represent the total amount of benefits enjoyed by all the
population which benefits from risk reduction.
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One approach for estimating the willingness to pay to avoid a
statistical death is to observe human behaviour in risky situations.
Most empirical estimates, which have been reviewed by Bailey (1980),
examine wage differentials among occupations with varying degrees of
risk. One empirical estimate used individual choice with respect to
seat belt use. The values (1978 dollars) found in the former studies
(wage differentials) ranged from approximately $250,000 (Thaler and
Rosen 1976) to $5.0 million (Smith 1974). The value found in the
seat belt study was approximately $313,000 (Blomquist 1979).
Other approaches for estimating the value of human life include total
lifetime earnings, court awards, and surveys. Current economic
thinking questions these approaches on theoretical and empirical
grounds.
Neither the behavioural nor the survey approach captures the
willingness-to-pay of relatives or close friends. One study
(Needleman 1976) indicates that including others' willingness to pay
could increase the statistical value of life by 25-100%. Although
this study measured willingness to pay, it differed from the
behavioural approach in that it placed a value on a known human life.
The behavioural approach assigns a value to an improvement in safety
for each of a large number of individuals.
Review of the significant behavioural studies could provide high and
low limits for the range of values of a statistical death avoided.
Therefore, it is the approach recommended for valuing the effects
associated with LRTAP. However, no monetary estimates are possible
unless there is an agreed upon dose-response relationship.
7.3.5.2 Morbidity
The conceptually correct procedure for estimating the value of
reductions in morbidity is also what an individual must be paid to
accept additional risk. Individuals must be paid a certain amount to
accept lost time at work, or restricted activity days. A more
complete analysis would also ask what an individual must be paid if
he had to accept a career change as a result of an accident.
Unfortunately, there are few behavioural studies and surveys which
provide us with estimates of willingness to accept risk. In lieu of
this information, average daily earnings (not wage differentials by
occupation) for those in the labour force can be used as an empirical
value with the recognition that not all morbidity results in lost
earnings (paid sick leave and sickness on nonworking day). Their
earning measures do not reflect loss in productivity and the pain and
discomfort they suffer.
The value of changes in morbidity would partially follow the lower
bound estimates of Freeman (1979). Morbidity could be measured
either by work days lost or restricted activity days. The work days
lost measure applies only to people in the labour force. Restricted
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also be used.
7.3.6 Visibility
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activity days applies to all people of all ages and includes degrees
of illness and incapacitation which are not severe enough to result
in absence from work. Work days lost and restricted activity days
could respectively be valued at $40-$50 per day and $10-$20 per day
to provide a range, the latter being the (U.S.) average gross daily
earning in the private nonagriculture sector in 1980. •
Other measures of the value of health effects can be obtained from
changes in medical expenditures for health care. In addition, the
costs, (e.g., relocation) incurred to avoid unhealthy situations can •
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The conceptually correct procedure for valuing changes in visibility _
is to estimate the willingness-to-pay in each region (Rowe and •
Chestnut 1981). The demand curve for an individual site would relate
the number of days of satisfactory visibility to the price of these
days, assuming no changes in such things as income and tastes (Figure •
7-4). If the number of days of satisfactory visibility is OA, the •
value of visibility is the entire area under the demand curve,
because there is no expenditure for visibility. This assumes the •
initial level of visibility is maintained. •
Using a dose-response relationship specified by the effects group, a _
reduction in LRTA.P with improved visibility increases the number of •
days of satisfactory visibility. This results in a movement along •
the demand curve. The net economic benefit is the increase in
willingness-to-pay as measured by the entire area under the demand
curve.
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The problem with this theoretical approach for valuing visibility, as •
with many environmental goods, is due to its special status as a •
"public good." There are no markets for which prices and demand
curves can be directly obtained. Thus, imputed market and nonmarket
approaches are proposed as valuation techniques in this field. •
The imputed market approach (hedonic prices/demand analysis) uses
existing market data, in cases where the selection of a market good
may vary with visibility levels, (e.g., the choice of residential
location). This approach further assumes that the intensity of these
preferences is revealed by individuals' behaviour and their demand _
for associated market goods (e.g., how much more individuals pay for •
homes in neighbourhoods with clean air, and the degree to which
vacationers change their travel plans reveal how much they value
visibility). Technical measures of pollution concentrations or •
visibility levels must be reasonable representations of the •
environmental attributes that individuals value. These measures must
be able to be used to identify that part of an individual's behaviour •
attributable to the component of environmental quality being •
studied.
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$
0
B
days of good
visibility
Figure 7-4. Change in demand due to visibility improvement.
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7.3.7 Summary
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The nonmarket approaches (e.g., Bidding Game and Rank Ordering)
attempt to elicit values through surveys of how respondents think •
they would react to a proposed visibility change. In contrast to the •
market approaches, nonmarket approaches do not attempt to infer
values of a component of environmental quality from observation of
individuals' actual behaviour in response to a change in •
environmental quality. Instead, individuals are asked to predict how •
they would respond to a change in environmental quality. Bias of
values determined by this method may be due to the level of •
information conveyed to respondents. This approach presupposes that |
a particular change in environmental quality can be described to the
respondents, usually with photographs and verbal descriptions, in a _
way that corresponds to what their perceptions of the actual •
experience would be. For example, it is assumed that a photograph of
the Grand Canyon obscured by pollution will elicit a response that
corresponds to what the response to the actual situation would be. •
This type of approach also assumes that individuals are capable and •
willing to predict their response behaviour to a hypothetical
situation that they may not have ever experienced. •
It is recommended that a review of empirical studies on the value of
visibility be undertaken to provide a range of values for various _
regions, and for urban and rural areas. •
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The following table provides a summary of methods and their
applicability to the various LRTAP affected receptor categories. •
The 'X' denotes that the method can be used, whereas 5C denotes the •
method which is recommended as most appropriate. The methods capture
only the primary values and that regional econometric analysis is _
necessary to draw out the secondary economic effects (e.g., jobs and •
income) in a given sector and in a specific area. •
7.4 QUALIFICATIONS, CONCLUSIONS AND RECOMMENDATIONS |
This section has provided a review of methods which can be employed «
to determine the primary economic benefits of LRTAP reduction on •
specific receptor categories, as well as the secondary economic
effects.
7.4.1 Qualifications
Although numerous limitations and qualifications have been noted, |
with respect to specific methods or issues, there are three
significant qualifications which are relevant to LRTAP-related _
environmental effects: •
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TABLE 7-2. SUMMARY OF METHODS
Market
Factor Substi- Travel
Income tutes Cost
A. Aquatic
1. Sports
Fishery X
2. Commercial
Fishery Xa
3. Ecosystem
B. Terrestrial
1 . Crops Xa
2. Forests Xa
3. Ecosystem
C. Buildings,
Structures
1 . Materials Xa
2. Historic
D. Water Systems Xa
E. Health
1. Morbidity
2. Mortality
F. Visibility
a X denotes the method is recommended as
7-27
Imputed Market Nonmarket
Property Observed Survey
Value Behaviour
Xa X
Xa
X
X
Xa
X
Xa
Xa
Xa
X Xa
the most appropriate.
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7.4.1.2 Inclusion of All Values
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1) there is a lack of dose-response relationship information;
2) there is difficulty in capturing all benefit values; and
3) there needs to be an evaluation of irreversibilities and •
the all or nothing feature. •
7.4.1.1 Dose-Response Relationship •
The need for data from the Effects Groups for the various receptor
categories has been stressed at several points. Although some data •
are available, a clear statement is needed of changes in output •
(e.g., water availability with fish populations) as related to LRTAP ™
effects (i.e., changes in pH). This must be further extrapolated
over geographical areas and over the short and long term to derive •
estimates of total quantity changes (Table 7-3). •
In the absence of these data, meaningful benefit estimates are •
impossible. Changes in producer cost would provide an alternative •
estimate of benefits of LRTAP control with yield and catch held
constant.
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A second concern is the extent to which the methods recommended will •
fully capture the value of the benefits. Some methods can provide J
only a partial measure, since they cannot capture option and legacy
values. Although their exclusion results in an underestimate, M
determination of the actual size of this underestimate is difficult. •
Some economists think the underestimates are large in situations
dealing with unique assets, or major changes in an entire
geographical region (e.g., New England). The matter is further •
complicated by the issue of property rights, discussed under equity ™
consideration. Thus, one should be cautious in assuming that any
benefit figure is a reflection of the full value to society. This •
may be less of a concern where measurable values are sufficient to |
indicate the desired choices.
7.4.1.3 Irreversibilities and the All or Nothing Feature •
There are additional limitations to conventional economic analysis.
The physical dose-response relationships with respect to LRTAP I
deposition may be irreversible and the rate of damage may not be •
monotonically related to deposition. This is called the all or
nothing feature or nonconvexities. •
First, once a certain level of damage has occurred, reduction in
LRTAP may not result in an improvement in environmental quality.
Hence, the effects of LRTAP may be irreversible (i.e., certain •
species may never be restored). If so, current market or inferred •
prices will substantially understate the value of these resources to
society. From the perspective of benefit valuation, it is imperative •
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TABLE 7-3. SUMMARY OF PHYSICAL SCIENCE DATA NEEDED FOR BENEFIT
EVALUATION
Inventory
Dose-Response
A. Aquatic
1. Sports
Fishery
water availability by
susceptibility, geographical
area and species type
2. Commercial water availability by
Fishery susceptibility, geographical
area and species type
3. Ecosystem
B. Terrestrial
species diversity, numbers
1. Agricultural crop pattern by geographical
Crops area
2. Forests
3. Ecosystem
C. Buildings,
Structures
1. Material
2. Historic
D. Water Systems
E. Health
1. Morbidity
2. Mortality
F. Visibility
cover type, age, stocking
and size by geographical
area
species diversity, numbers
geographical distribution
by type of material and by
use
geographical distribution
by type of material
geographical distribution
of systems on susceptible
water bodies
population
population
population
change in fish
population with
varying deposition
levels
change in fish
population with
varying deposition
levels
changes in species
diversity and numbers
change in marketable
yield with varying
deposition levels
change in marketable
yield with varying
deposition levels
changes in species
diversity and numbers
deterioration rate
as a function of
total sulphur
deterioration rate
as a function of
total sulphur
change in lake/stream
intake pH
sickness per
deaths per yg/m3
change in km of
visibility per
3 S042~
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7-30
that option, existence, and legacy values be included in the
estimates.
property rights.
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Second, after a certain level of pollution, the rate of damages
declines (Crocker and Forster 1981). Such would be the case where
most of the fish are gone and further pollution has little or no •
impact. This feature of nonconvexity would suggest that benefits of |
LRTAP control are much higher, increasing at a faster rate in a
relatively unpolluted environment. Once the rate of damages starts •
to decline, the benefits of abatement would be commensurately lower. •
The limitation of this suggestion is that it is based on the adverse
effect on a few species. If acidic deposition affects numerous
species, any mitigation effort even at high levels of pollution could •
show significant benefits. •
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7.4.2 Conclusions and Recommendations
This paper has attempted to provide an overview of techniques of .
deriving the economic value of benefits associated with LRTAP •
abatement. There is a large body of economic literature which deals ™
more thoroughly with the intricacies of the theory and is replete
with numerous empirical studies. However, many of the latter have H
not dealt specifically with the effects of LRTAP. |
Four conclusions arise from the material presented here. •
1. There are several techniques which can be applied to determine
the primary economic benefits associated with a particular
activity category. The values are underestimated since they •
fail to include option and legacy values for some effects. B
However, the lack of data on dose-response relationships limits
the application of these techniques at this time. •
2. Even in the absence of dose-response data, a variation in the
factor income approach is available to estimate the benefits of _
changes due to reduced LRTAP deposition. This approach provides I
benefit estimates on the basis of the differences due to various ™
levels of LRTAP in production costs for a given level of output
and could be applied to commercial fisheries, agriculture, H
forestry, and buildings and structures. •
3. The value of the benefits can be further estimated for specific •
economic sectors, and hence regions, to derive an estimate of •
the impacts in various geographical areas.
4. It is evident that more economic research is required. Economic •
techniques have yet to be rigorously tested in some sectors, ^
(e.g., historical value) and are limited in their treatment of
option and legacy values and in dealing with the issue of •
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The following recommendations can be made:
1. Dose-response data are available for the aquatic receptor and
the geographically - specific studies now being undertaken tend
to ignore substitution among fishing sites. Therefore, the
participation model should be applied on a U.S./Canada basis to
sports fishing to determine the value of primary benefits due to
LRTAP reduction.
2. A regional economic analysis should be undertaken to derive the
secondary value of the recreation and tourism sector in areas of
the U.S. and Canada affected by LRTAP (e.g., Adirondacks and
Muskoka-Haliburton).
3. To develop benefit estimates for LRTAP reduction for commercial
fisheries, agriculture, forestry, and buildings and structures,
a variation on the standard factor income approach should be
used. Here the differential in the cost of producing a given
level of output is determined.
4. Further research should be undertaken to determine the most
appropriate value for changes in morbidity.
5. Further research needs to be initiated to apply the survey
(contingent market) methodologies to the derivation of primary
benefit values of visibility in the eastern U.S. and Canada and
to historical sites, because of the lack of information about
these values.
6. Further work needs to be undertaken with respect to the issues
relating to property rights. These are an important part of the
distributional aspect of the long range transport of
pollutants.
7. The relationship between activity and other (option and legacy)
values for the various receptor categories should be further
investigated in order to derive a sense of the underestimate of
the total benefits due to the omission of the latter values.
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7.5 REFERENCES
Baltimore, MD.: Johns Hopkins University Press for Resources
for the Future.
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Bailey, M. J. 1980. Reducing risks to^ life. American Enterprise I
Institute, Washington, DC. •
Bender, L.D. 1975. Predicting employment in four regions of the •
western United States. Technical Bulletin No. 1529, Economic |
Research Service, U.S. Department of Agriculture, in
cooperation with Montana State University Agricultural _
Experiment Station. I
Blomquist, G. 1979. Value of life saving: implications of
consumption activity. Journal of Political Economy (87). I
Brookshire, D.B.; Ives, B.; and Schulze, W.D. 1976. The valuation of
aesthetic preferences. JEEM 325-46. •
Canadian Treasury Board. 1976. Benefit-cost analysis guide.
Canadian Treasury Board Planning Branch, Information Canada, _
Ottawa, Ont. •
Conoposk, J.V. 1978. A data-pooling approach to estimate employment
multipliers for small regional economies. Technical Bulletin •
No. 1583, Economics, Statistics, and Cooperatives Service, |j
U.S. Department of Agriculture, in cooperation with the U.S.
Environmental Protection Agency. •
Crocker, T.D., and Forster, B.A. 1981. Decision problems in the *
control of acid precipitation: nonconvexities and
irreversibilities. J. Air Pollut. Control Assoc. 31(1): B
31-37. 1
Crocker, T.D.; Tschirhart, J.T.; Adams, R.N.; and Forster, B. 1981. •
Methods development for assessing acid precipitation control B
benefits. U.S. Environmental Protection Agency, draft
report. _
Eckstein, 0. 1961. A survey of public expenditure criteria. In ™
Public finance: needs, sources, and utilization.
Universities-National Bureau Committee on Economic Research, B
Princeton, NJ.: Princeton University Press. B
Freeman, A.M. III. 1969. Project design and evaluation with multiple mm
objectives. In The analysis and evaluation of public I
expenditures: the PPB system, subcommittee on Economy in
Government, U.S. Congress Joint Economic Committee,
Washington, DC. I
1979. The benefits of environmental improvement.
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Haveman, R.H., and Weisbrod, B.A. 1975. The concept of benefits in
cost-benefit analysis: with emphasis on water pollution
control activities. In Cost-benefit analysis and water
pollution policy, eds. H.M. Peskin and E.P. Seskin. Urban
Institute, Washington, DC.
Maler, K.G., and Wyzgon, R.E. 1976. Economic measurement of
environmental damage: a technical handbook. Paris: OECD.
McKean, R.N. 1958. Efficiency in government through systems
analysis. New York: John Wiley and Sons.
Mishan, E.J. 1971. Cost-benefit analysis. New York: Praeger.
Needleman L. 1976. Valuing other people's lives. Manchester School
of Economic and Social Studies 44: 309-342.
Ridker, R. 1967. Economic costs of air pollution. New York:
Praeger.
Ridker, R. , and Henning, J. 1967. The determinants of residential
property values with special reference to air pollution.
Review of Economics and Statistics 49: 246-257.
Rowe, R.D., and Chestnut, L.G. 1981. Visibility benefits assessment
guidebook. Interim Report to U.S. Environmental Protection
Agency, Contract Number 68-02-3528.
Russell, C.S. 1981. Measuring the damages of acid precipitation
deposition; recreational fishing. Memo to USEPA.
Steiner, P.O. 1977. The public sector and the public interest. In
Public expenditure and policy analysis, 2nd edition, eds. R.H.
Haveman and J. Margolis. Chicago: Rand-McNally.
Thaler, R., and Rosen, S. 1976. The value of saving a life. In
Household production and consumption, ed. N.E. Terleckyj.
National Bureau of Economic Research, New York, NY.
U.S. Water Resources Council. 1980. Proposed rules; principles,
standards and procedures for planning water and related land
resources. Federal Register Vol. 45 April 14.
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APPENDIX - REVIEW OF RELEVANT ECONOMIC CONCEPTS
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Two basic economic concepts which provide a measure of changes in •
social welfare (satisfaction) are consumer's surplus and producer's ™
surplus. Changes in these measures are regarded as being the most
theoretically relevant indicators of social welfare loss or gain, •
resulting from specific activities or events. |
A. 1 Consumer's Surplus I
Traditional economic theory presumes that the individual consumer is
the best judge of his own personal well-being or utility given I
currently available information. If an individual is made better •
off, other things being equal, then his social well-being or welfare
is increased. •
The individual consumer allocates his money income across the various
commodities in such a fashion that he maximizes his welfare or •
utility. In general, his desired purchase of a commodity will depend I
upon his tastes, the prices of all goods, and his income.
The demand curve graphically represents the relationship between the •
desired purchase of a commodity and its price (or the willingness- •
to-pay). For each additional unit, the consumer is willing to pay
less than for the previous unit. Hence, the curve slopes down to the
right. This is called the ordinary demand curve or the Marshallian
demand curve. If we assume that more of the good will be purchased
at lower prices if prices fall (a "normal" good), then a consumer's _
ordinary demand curve is represented in Figure 7-5. •
A point on the demand curve is the maximum price that the individual
would be willing to pay for a specified amount of good, and is noted I
by an ordered pair (x, p). Alternatively, for a given price p, x •
represents the most the consumer would willingly purchase. A maximum
price exists for every potential consumption level for the good, and
is given as the relevant p-point on the demand curve.
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Suppose that the commodity sells in the market for PI and the con- _
sumer purchases Xj. The consumer's expenditure on x is Pj X^ B
(price times quantity). The triangular area denoted PjAB, which •
lies above the expenditure rectangle OPjBXj is what economists
call the Consumer's Surplus. It is a surplus, since it represents a I
saving to the individual in terms of what he would have been prepared |
to pay for levels of consumption smaller than X, as shown by the
associated prices on the demand curve. Instead of paying the maximum M
price for each level of commodity x, the consumer pays Pj for all B
units. If all of the savings are added up, then we obtain the area
P^AB. Since the price is given in money units, consumer surplus is
a monetary measure. The area OP^BX^, plus the area PjAB B
(consumer expenditure plus consumer surplus), is a measure of the B
gross benefits to the individual of consuming Xj units. Consumer's
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o n>
Q.
1 If
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O)
1 li
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Pi
\
*~ \ consumer surplus
^V
< expenditures
X
X1
Figure 7-5. Measure of consumer surplus.
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1
P1
P2
0
V
\ox
x1 x2
Figure 7-6. Change in consumer surplus.
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surplus is the net benefit to the individual consuming x° units,
that is, total willingness-to-pay minus expenditure.
Changes in the value of consumer's surplus are a measure of the •
welfare associated with the activity which caused the change.
Consider the case in which the price of the commodity falls from P^ H
to ?2 due to an environmental improvement as illustrated in Figure |
7-6. At price ?2, the consumer can both purchase more of the
commodity, and pay less per unit. In addition, his consumer's _
surplus is increased by the trapezoidal area P2PiEB. This I
geometric area represents a monetary measure of the welfare effect ™
associated with the price change.
Consumer's surplus can be estimated for an individual, using observed •
price and quantity data. However, instead of estimating individual
demand curves, economists use aggregated data and estimate market •
demand curves. Market demand curves are obtained by aggregating •
individual demand curves, that is, adding up horizontally (along the
quantity axis). This implies that tastes and preferences can be _
aggregated across individuals. If, however, individuals have •
different incomes, or if the distribution of income is altered ™
significantly, then aggregations can lead to biases in estimates.
Simple linear summation of these demand curves is inadequate. Then, H
one must resort to Engel curves. I
There are two alternative monetary measures of the effects of a price •
change, known as equivalent and compensating variation (denoted as EV •
and CV respectively), which can be translated into a change in
income. Under the circumstances where a decrease in LRTAP effects
results in a price decrease, EV and CV can be defined as follows: I
Compensating Variation is the change in income which, given the price
decrease, maintains the consumer's original utility. CV is equal to
the income which would be withdrawn to offset the price decrease.
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Equivalent Variation is the change in income which, given the _
original price, would leave the consumer's satisfaction or utility I
unchanged if price decreases. An increase in income equal to EV
would be given to the consumer to maintain welfare.
While EV and CV are technically the more correct measures of welfare |
change, they are difficult to estimate. The value of consumer
surplus, which is closely related to CV and EV, is easier to measure •
and is therefore recommended for this analysis. I
Public Goods —
In the above discussion, we assumed that commodity x was traded in an •
organized market at a nonzero price. The impact of LRTAP on the
price of a particular commodity was subsequently considered. This •
scenario is, of course, an over simplification. Now let us consider |
a certain commodity which is a "public good" such as an environmental
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commodity (e.g., clean air). This type of commodity is not traded in
an organized market but it does have a value to society. The levels
of this environmental commodity are assumed to be outside the control
of the individual. Hence, the overall level of consumer welfare
depends not only upon prices of market commodities and money income,
but also upon the level of environmental commodities he consumes.
Where there are no markets and hence no prices, it is difficult to
derive a demand curve. Demand is derived through other means as
discussed in Sections 2 and 3. Again, the measure of consumer's
surplus is the area under the demand curve. In this case, consumer's
surplus is the total area since the price is zero. The change in
consumer surplus would be measured by the change in quantity if, for
example, visibility increases due to reduced LRTAP (Figure 7-7).
While willingness-to-pay is used to determine demand curves, this is
not the real test of the value of visibility. However, it is an
easier measure. The change in quantity has affected consumer
utility, and the consumer effectively enjoys an increase in income.
We can therefore obtain a monetary measure of the welfare change, by
considering the change in income which will have the same impact as
the change in environmental quality. Here there are two measures —
compensating and equivalent surplus (denoted as CS and ES), depending
upon which welfare position is used for the initial starting point
for comparison.
Compensating surplus is the change in income which results in the
same level of utility, given the change in quantity (Figure 7-8).
Equivalent surplus is a change in income which produces a change in
utility equal to the change in quantity, at the original quantity
level.
A.2 Producer's Surplus
The discussion thus far has been concerned with consumer's surplus
as one measure of economic welfare. It is possible to define an
analogous concept for producers in the economy. This is called
producer's surplus. The concept of consumer's surplus is defined
with respect to the consumer's demand curve. Producer's surplus is
defined with respect to the producer's supply curve of the relevant
commodity. Figure 7-9 presents a supply curve, which presumes that
more of the output will be produced as price rises. Higher prices
are required to cover increased production costs at higher output
levels.
In Figure 7-9 a point (Xj, Pj) on the supply curve can be given
two interpretations. For a given price P^, the output Xj is the
largest that the firm is prepared to supply at that price. For a
given output Xj, the price PI is the minimum price that the firm
will accept for supplying X^. In the market all units sell for the
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7-38
$
0
A B days of good
visibility
Figure 7-7. Change in demand due to visibility improvement,
Income
M
Figure 7-8. Compensating and equivalent surplus.
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7-39
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x-i
Figure 7-9. Producer's surplus.
A
E
0
Figure 7-10. Change in producer's surplus due to change in supply.
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7-40
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same price and the producer gains on all unit levels which are lower
than the total sale. This is because the market price exceeds the
minimum the producer needs. This gain is called producer's surplus, •
and is represented by the area AP^B, the area above the supply |
curve bounded by the market price (Figure 7-9).
Changes in the value of this producer's surplus are interpreted as a •
measure of welfare change. This change causes a supply shift from S
to Sj, due, for example, to an increase in crop yields from reduced
LRTAP (Figure 7-10). •
The area ABCE represents the welfare gain to the producer caused by a
shift in the supply curve from S to S^. The minimum price required •
to supply each level of output is now lower, and is everywhere £
further from the market price received by the producer.
Changes in net social welfare caused by LRTAP effects on marketable I
commodities can be determined by examining the net change in
consumer's and producer's surpluses. Suppose, for example, that the
reduction of LRTAP deposition results in an increased supply of some I
product. The supply curve has then shifted to the right from S to m
Sj_, while the demand curve for the product remains stationary at D,
(Figure 7-11). •
The area EP2C is the new producer's surplus, caused by the price
fall due to the supply increase, compared to AP^B at the original _
supply and price levels. Producer's surplus changes for two reasons. •
The producer's surplus is increased by EAFC as a result of increased ™
production at lower cost, with a given market price. Producer's
surplus decreases by P2P^BF a result of market pressures •
decreasing output prices and stimulating production. The net change |
in surplus is therefore ABCE.
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7-41
Pl
P
Figure 7-11. Hypothetical change in producer's surplus due to
reduction in LRTAP deposition.
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SECTION 8
NATURAL AND MATERIAL RESOURCES INVENTORY
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8-1
SECTION 8
NATURAL AND MATERIAL RESOURCES INVENTORY
8.1 INTRODUCTION
The objective of an economic evaluation of acidic deposition is to
estimate the value of reduced adverse effects achieved by a given
reduction in emissions. This objective is pursued in six steps:
1. Inventory: Identify the totality of the resource of interest.
2. Sensitivity: Divide the resource inventory into sensitivity
classes.
3. Exposure: Subdivide each sensitivity class by the severity of
some indicator of deposition.
4. Response: Measure the adverse response expected absent of
mitigative measures, for each sensitivity-exposure subdivision.
5. Mitigation: Determine reduction in deposition resulting from
mitigation measures and calculate the fraction of the adverse
effects estimated in (4) that would be reduced.
6. Valuation: Estimate the value of the adverse effects reduced.
The long-range transport of air pollutants (LRTAP) inventory of
resources potentially at risk includes aquatic, terrestrial, and
man-made resources. In all cases, the inventories now available are
incomplete and generally lacking in the detail needed for a
benefit/cost evaluation. The aquatic inventory is limited to large
streams and lakes, and does not include potentially affected fish
populations or the many plants, insects and animals living in or
adjacent to water bodies. The terrestrial ecosystem consists of two
major components, agriculture and forestry. The inventory of each
will be conducted separately. Only major crop values and production
are surveyed for the agricultural inventory. The forest inventory
differentiates only between major forest types and does not include
any information on shrubs and grasses. The materials inventory is
far. from complete in that it does not include common construction
materials, such as galvanized steel and chain link fence. It lacks
detail in describing historic places, landmarks and parks, and is the
least comprehensive of the three categories.
This LRTAP inventory of resources potentially at risk does not
include all natural and man-made resources in eastern North America.
Wherever data are available, the inventory is geographically
selective by two important criteria: (1) sensitivity, and (2)
deposition regime.
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8-2
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The first criterion is the sensitivity of the various natural and
man-made systems to acidic deposition. (The concept of sensitivity
is explained in more detail in sections 3.5 and 4.5 of this report). •
The sensitivity of aquatic ecosystems is a function of soil |
characteristics, bedrock geology, topography, and alkalinity of the
receiving waters. The sensitivity of terrestrial ecosystems is a M
function of soil characteristics and management practices and bedrock B
geology. It should be noted that even if a forest ecosystem is not
in a sensitive area, its foliar system may still be affected by
acidic deposition. The sensitivity of man-made structures is a B
function of the specific material and the mitigation measures B
undertaken by man. For example, the sensitivity of metals is a
function of their composition and of the surface platings or coatings •
of corrosion resistant materials. Calcareous stone and masonry are |
sensitive materials unless protected.
The second criterion is the intensity of acidic deposition. Wet B
sulphate deposition is used herein as an indicator because data are
available and because wet sulphate deposition is clearly an important
contribution to overall acidification. Other factors, (e.g., dry B
deposition, nitrates, and seasonability of deposition), are known to |
affect the acidification potential of deposition, but an indicator
which combines all of those factors is not yet available. It is m
known that ambient sulphur dioxide concentration is a more B
appropriate indicator of the potential damage to materials than wet
sulphate, so SC>2 is used in place of sulphate when considering —
materials. Wet sulphate deposition is divided into three ranges as I
shown in Figures 8-la and 8-lb: low (10-20 kg/ha.yr), moderate B
(20-40 kg/ha.yr), and high (greater than 40 kg/ha.yr).
deposition intensity to define resources potentially at risk is best
explained by a simple graphic (Figure 8-2). Each data category ^
(e.g., resource distribution, sensitivity and deposition) constitutes B
one set. Any overlap of the three sets defines the resource *
potentially at risk. Thus, the inventories provide information on
the quantity and nature of resources within each of the three I
deposition zones. In the case of aquatic resources, this is |
supplemented by estimates of the potential of the soils and bedrock
to reduce (or buffer) acidity. •
The estimates of resources at risk presented in the following
sections are based on steps 1 to 3, (i.e., inventory, sensitivity, —
and exposure; page 8-1) and are illustrated in Figure 8-2. Steps 4 I
to 6 (i.e., response, mitigation, valuation), as well as better data ™
for steps 1 to 3, will further reduce the amount of the resource of
interest in evaluating an emission reduction measure. It should be •
clear from the other sections in this report that our ability to B
perform steps 4 to 6 is limited at present. Therefore, the estimates
below should not be interpreted as representing the value attri-
butable to a deposition control measure, but rather as categories of
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8-3
0-10 KG/HR
10-20 KG/HR
20-40 KG/HR
>MO KG/HR
Figure 8-la. Annual sulphate deposition regime for eastern United
States, based on NADP data covering April 1979 to
March 1980.
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8-4
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8-5
SPECIFIC RESOURCE
SENSITIVITY
RESOURCE INVENTORY
DEPOSITION PATTERN
Figure 8-2. Conceptual scheme for identifying resources
potentially at risk.
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8-6
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resources potentially at risk from which actual damages remain to be
derived.
The resource inventory has drawn on data from a variety of sources. |
Although none of it was compiled specificially for acidic deposition,
the best estimates have been made using this information compiled for •
different purposes, in different ways and for various years. •
Attempts were made to ensure that the U.S. and Canadian inventories
are reasonably comparable. Despite the minor differences it is
believed that the inventory presented here represents the best data •
available. Additional data collection will be necessary to improve I
the inventory both in coverage and specificity (e.g., tree species).
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8.2 AQUATIC ECOSYSTEM
For the purpose of defining potential resources at risk, the aquatic I
ecosystem identified here pertains only to lake and stream area
measures. A census of surface water resources within each
combination of sensitivity/deposition regimes has been taken to M
provide quantitative estimates of the total area of surface waters •
(i.e., lakes and streams) potentially at risk.
8.2.1 U.S. Aquatic Resources
The Work Group took as its starting point for an inventory of surface •
water areas the Geoecology Data Base maintained by Oak Ridge National ™
Laboratory (ORNL). The surface water inventory in the Geoecology
Data Base includes all lakes greater than 2 acres and permanent
streams. The primary advantages of using the inventory in the
Geoecology Data Base are the completeness of the surface water
inventory and that ORNL prepared the map of sensitive areas for the M
Aquatics Subgroup (Figure 3-10; Olson et al. 1982). The primary •
disadvantages of using the Geoecology Data Base are the absence of '
data on surface water chemistry (i.e., alkalinity) and the inability
to discriminate among various sizes of lakes and streams. The B
inventory includes several large lakes and streams which even in •
sensitive areas would probably not be adversely affected by acidic
deposition. •
The Work Group limited its inventory effort to 38 states; those east
of the 100° meridian. The surface water area in all counties with _
50% of the land area in urban and agriculture uses was assigned to a •
special category rather than one of the three sensitivity categories. ™
Surface water in this category were assumed to be more adversely
affected by urban and agricultural activities than by acidic •
deposition. The remaining surface water area was assigned to one of |
four deposition categories. The disaggregated results of the
classification are included in Appendix Tables 8-1 to 8-3. M
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8-7
Approximately 25% of the U.S. surface water area is located in areas
with limited (low and moderate) potential to reduce acidity and of
deposition greater than 20 kg S042-/ha.yr (Table 8-1). Only
10% are located in areas with the most limited (low only) potential
to reduce acidity and of deposition greater than 20 kg SC>42-/ha.yr.
The actual surface water area would be more limited if data were
available on surface water chemistry (i.e., alkalinity). Additional
refinements of the inventory should include data on this variable as
well as more accurate measurements of surface water area.
Although the aggregate 38-state data show that approximately 25% of
the U.S. surface water area is potentially at risk, data disaggre-
gated at the state level show a much higher percentage in some states
(Table 8-2). All of the New England states have at least 70% or more
of the surface waters potentially at risk. Three mid-Atlantic states
(Maryland, New Jersey, and New York) and three southern stated
(Georgia, North Carolina, and Virginia) have at least 50% or more of
their surface water potentially at risk. Most of the states in the
mid-west, west and southwest have very limited, if any, surface water
potentially at risk except for Michigan and Arkansas. In some states
with low potential to reduce acidity and numerous lakes, such as
Wisconsin and Minnesota, the annual sulphate deposition loading is
less than 20 kg/ha.yr, so the surface water area is not considered at
risk. In all cases, these estimates of surface water potentially at
risk will be reduced to some degree when data on stream chemistry are
available.
8.2.2 Canadian Aquatic Resources
The basis for the inventory was provided by the map indicating the
potential of soils and bedrock to reduce the acidity of atmospheric
deposition (Figure 3-9; Lucas and Cowell 1982). This was overlaid
with the map of sulphate deposition (Figure 8-lb). Finally, data on
the proportion of surface water area for each province was combined
with the deposition/acidity reduction capability information to
derive estimates of the total area of surface waters at risk.
The data on surface waters were drawn from two main sources. In
Ontario, detailed lake counts and measurements (Cox 1978) provided
data on a watershed basis. For Quebec and the Maritimes, the
Ecodistrict Data Base developed by Environment Canada (1981a,b) was
utilized. The data presented here provide an estimated ratio of
water to land for each Ecodistrict. No data were available for
Newfoundland and Labrador. Two other serious omissions of this
inventory are a lack of information on lake alkalinity and data on
specific aquatic biota associated with the various deposition regimes
on a provincial basis.
Table 8-3 provides a provincial summary of the aquatic resources at
risk based on surface water sensitivity (as estimated by the
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8-8
TABLE 8-1. SUMMARY OF EASTERN U.S. SURFACE WATER AREA (km2)
CROSS CLASSIFICATION BY SENSITIVITY AND DEPOSITION
(USDA 1971, 1978a)
Sulphate Urban and
Deposition Agricultural
(kg/ha.yr) Area
AREA km2 (Percent of Total)
Potential to Reduce Acidity
Low3
Moderate
High
Totalb
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0-10 2,290(2) 10(<1) 1,240(1)
10-20 17,240(12) 5,410(4) 11,470(8)
20-40 11,120(8) 13,940(10) 19,970(14)
40 5,070(3) 210(<1) 2,190(1)
TOTALb 35,720(25) 19,570(15) 34,870(25)
3,950(3) 7,490(5)
15,740(10) 49,860(35)
27,400(19) 72,430(50)
7,620(5) 15,090(10)
54,710(35) 144,870(100)
a A low potential to reduce acidity is interpreted as a high
sensitivity.
k Rounded to the nearest 5%.
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8-9
TABLE 8-2. SURFACE WATER AREA WITH LOW AND MODERATE POTENTIAL TO
REDUCE ACIDITY RECEIVING GREATER THAN 20 kg/ha.yr
SULPHATE DEPOSITION (USDA 1971, 1978a)
km2 (percent of State Total)
Low Potential to
Reduce Acidity
Moderate Potential
to Reduce Acidity
REGION I
Connecticut
Maine
Massachusetts
New Hampshire
Rhode Island
Vermont
REGION II
New Jersey
New York
REGION III
Delaware
Dist. of Columbia
Maryland
Pennsylvania
Virginia
W. Virginia
REGION IV
Alabama
Florida
Kentucky
Georgia
Mississippi
North Carolina
South Carolina
Tennessee
REGION V
Illinois
Indiana
Michigan
Minnesota
Ohio
Wisconsin
REGION VI
Arkansas
Louisiana
Oklahoma
Texas
340
6,010
970
860
340
720
150
1,930
0
0
20
40
70
120
10
0
30
250
60
310
0
90
0
10
1,690
0
0
0
130
0
0
0
( 70)
(100)
( 90)
(100)
( 75)
( 70)
( 15)
( 40)
( 0)
( 0)
( <5)
( <5)
( <5)
( 25)
( <5)
( 0)
( <5)
( 10)
( <5)
( 5)
( 0)
( <5)
( 0)
( <5)
( 40)
( 0)
( 0)
( 0)
( 5)
( 0)
( 0)
( 0)
0
0
0
0
0
0
660
1,250
0
0
1,020
550
1,660
20
1,280
410
400
1,660
810
9,010
640
110
0
40
0
0
0
0
1,350
0
1,090
10
( 0)
( 0)
( 0)
( 0)
( 0)
( 0)
(75)
(20)
( 0)
( 0)
(50)
(35)
(55)
( 5)
(35)
( 5)
(15)
(55)
(40)
(85)
(25)
(<5)
( 0)
(<5)
( 0)
( 0)
( 0)
( 0)
(35)
( 0)
(25)
(<5)
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8-10
TABLE 8-2. CONTINUED
km2 (Percent of State Total)
REGION VII
Iowa
Kansas
Missouri
Nebraska
REGION VIII
North Dakota
South Dakota
Low Potential to
Reduce Acidity
0 ( 0)
0 ( 0)
0 ( 0)
0 ( 0)
0 ( 0)
0 ( 0)
Moderate Potential
to Reduce Acidity
0 ( 0)
0 ( 0)
0 ( 0)
0 ( 0)
0
0
( 0)
( 0)
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8-11
TABLE 8-3. SUMMARY OF SURFACE WATER AREA (km2) IN EASTERN CANADA
CROSS CLASSIFICATION BY SENSITIVITY AND DEPOSITION
Ontario
Quebec
Maritimes
TOTAL
Sulphate
Deposition
(kg/ha. yr)
10-20
20-40
>40
10-20
20-40
>40
10-20
20-40
>40
AREA km2
Potential
Lowb
11,254(12)
8,452(9)
408(<1)
20,474(22)
10,137(11)
-
-
8,719(9)
-
59,444(63)
(Percent of
Total)3
to Reduce Acidity
Moderate
4,142(4)
1,890(2)
98(<1)
2,532(3)
730(<1)
456(<1)
-
13,447(14)
-
23,295(25)
High
1,672(2)
2,120(2)
408(<1)
2,972(3)
3,006(3)
252(<1)
-
1,482(2)
-
11,912(13)
Total
17,068(18)
12,462(13)
914(1)
25,978(27)
13,873(15)
708(<1)
-
23,648(25)
-
94,651(100)
a Total surface water area receiving more than 10 kg/ha.yr sulphate
deposition.
" A low potential to reduce acidity is interpreted as a high sensitivity,
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8-12
8.3.1 U.S. Agricultural Resources
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potential of surrounding soils and bedrock to reduce acidity) and
sulphate deposition.
Of the total estimated surface water area of 51,605 km2 located in |
regions sustaining over 20 kg/ha.yr of sulphate deposition,
44,337 km2 (86%) are in areas with either low or moderate potential M
to reduce acidity. More than half of this (27,716 km2) are in •
areas of low potential alone. Within the moderate and high deposi-
tion zones, the majority of surface water is receiving between 20 and
40 kg/ha.yr of sulphate; only 1.9% (962 km2) both receive more than •
40 kg/ha.yr sulphate and have low or moderate potential to reduce •
acidity.
The provincial breakdown indicates that 94% (22,166 km2) of the |
surface water surveyed (i.e., receiving more than 10 kg/ha.yr
sulphate) in the Maritimes both receive at least 20 kg/ha.yr of •
sulphate and have a low or moderate potential to reduce acidity. •
Although Quebec has the greatest total surface water area
(40,559 km2) only 28% (11,323 km2) are within areas of low and
moderate potential to reduce acidity receiving more than 20 kg/ha.yr M
sulphate. Thirty-six percent (10,848 km2) of surface water •
surveyed in Ontario are in a moderate or high deposition zone
combined with a low or moderate potential to reduce acidity. •
1
8.3 AGRICULTURAL RESOURCES —
The majority of crops listed in the inventory have been selected due
to their significance in terms of value or production. The six most
important crops are corn, soybeans, wheat, hay, tobacco and potatoes. B
This basic list has been supplemented by other crops which •
individually ranked high in the U.S. (cottonlint and sorghum) and
Canada (barley and vegetables). Maps which provided crop data on a •
county or census tract basis were overlaid with deposition informa- I
tion to provide the quantitative crop information. The inventory
presented here provides data on crop yields and values by state or
province for each of the three deposition zones. •
I
The growing of agricultural crops is a major economic industry in the
United States. Farms in the U.S. in 1978 produced over $64.9 m
billion worth of crops (USDA 1980). I
The U.S. Department of Agriculture each year publishes its estimates
of the previous three years crop statistics in Agricultural •
Statistics (USDA 1980). In addition to data on agricultural I
supplies, consumption, costs, and returns, this reference book lists
data on acreage, production, yield, and value of 99 crops grown in ft
the U.S. Of these crops, about 34 have been studied for their yield |
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8-13
response to acidic deposition. The inventory in this section was
limited to those crops of major economic importance among the 34
crops.
The major crops were identified by ranking all of the crops by their
1978 estimated value of production (USDA 1980). Table 8-4 presents
this ranking for the entire U.S. and each crop's cumulative percent-
age of the total. It was found that the top eight accounted for
almost 75% of the total value of all U.S. crops.
Of these eight crops, there are research studies on the effects of
acidic deposition on the yields of six. There are no effects data on
cottonlint and sorghum. Consequently, yield of only the six crops
studied is matched with sulphur deposition patterns.
Data on yield of these six crops in the states east of the 100°
meridian (38 states) is displayed in Appendix Tables 8-4 to 8-7.
The four tables describe yield for the six crops by deposition
pattern (i.e., 10-20 kg/ha.yr, 20-40 kg/ha.yr and greater than 40
kg/ha.yr) and total yield. Many states produce some of these crops
under all three deposition patterns.
The total yield of each crop under four deposition patterns shows
considerable variation (Table 8-5). Soybeans and tobacco are the
only crops with any significant proportion of their yield in areas
with sulphur deposition greater than 40 kg/ha.yr. For the remainder
of the crops, less than 15% of their total yield is grown in areas of
high deposition.
Although the aggregate 38-state data show that only 20% or more of
the yield of two of the six major crops receives sulphate deposition
greater than 40 kg/ha.yr, disaggregated data show that a higher
percentage of crops in some states receive a high rate of sulphate
deposition (Table 8-6). More than 50% of soybean yield in five
states and of tobacco yield in two states receive sulphate deposition
greater than 40 kg/ha.yr. In addition, a significant portion of the
six crops in some states receive a high rate of deposition. At least
50% of three crops in the states of Arkansas, Kentucky, Michigan,
Ohio and Tennessee receives 40 kg S04^~/ha.yr.
8.3.2 Canadian Agricultural Resources
Agriculture is an important economic activity for all provinces in
eastern Canada with most of the yield and value centred in Ontario
and Quebec. Data have been assembled from Statistics Canada and
provincial agriculture ministries to provide an overview of the
types, yields, and values of crops at risk within each of the three
identified deposition regimes. The crops of importance are primarily
grains, but data on certain vegetables are also included, although
they represent only about 1% of the total value of production.
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TABLE 8-4.
CROP
Corn
Soybeans
Hay
Wheat
Cottonlint
Tobacco
Sorghum
Potatoes
TOTAL
8-14
RANKING OF U.S. CROPS BY 1978 VALUE OF PRODUCTION
(USDA 1980)
1978 $ Value Percent Cumulative Production
(106) of Total Percent (Metric Tons 106)
15,900 24 24 177.2
12,500 19 43 50.9
6,600 10 53 126.6
5,400 8 61 48.9
3,000 5 66 2.4
2,700 4 70 .9
1,500 2 72 19.0
1,200 2 74 16.3
48,800
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8-15
TABLE 8-5. 1978 YIELD OF SIX CROPS IN 38 STATES BY DEPOSITION
REGIME (USDC 1979)
Metric Tons 106
(Percent of 38-State Total)*
Sulphate Deposition kg/ha. yr
CROP
Corn
Soybeans
Hay
Wheat
Tobacco
Potatoes
<10
7.2
(5)
.2
(0)
4.6
(5)
7.5
(25)
0
(0)
<.l
(0)
10-20
76.7
(45)
13.7
(30)
41.4
(45)
17.6
(60)
<.l
(0)
2.7
(50)
20-40
68.4
(40)
21.7
(45)
32.2
(35)
3.1
(10)
.7
(80)
2.3
(45)
> 40
16.7
(10)
11.3
(25)
13.8
(15)
2.2
(5)
.2
(20)
.3
(5)
Total
169.0
46.9
92.0
30.4
.9
5.3
To the nearest 5%.
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8-16
TABLE 8-6. SELECT U.S. AGRICULTURAL CROPS BY STATE RECEIVING
GREATER THAN 40 kg/ha.yr SULPHATE DEPOSITION
(USDC 1979)
Metric Tons 103
(Percent of State Total)3
REGION II
New York
REGION III
Maryland
Pennsylvania
W. Virginia
REGION IV
Kentucky
Mississippi
Tennessee
REGION V
Illinois
Indiana
Michigan
Ohio
REGION VI
Arkansas
Louisiana
Oklahoma
Texas
REGION VII
Missouri
Corn
404.7
(30)
7.9
( <5)
392.9
(15)
52.2
(35)
2,027.3
(70)
21.5
(15)
621.6
(60)
390.8
(<5)
1,107.6
(5)
2,268.1
(50)
8,565.1
(95)
26.5
(95)
3.2
(5)
.1
(<5)
312.9
(10)
430.1
(10)
Soybeans
0
(0)
0
(0)
2.7
(5)
0
(0)
839.5
(85)
805.7
(40)
1,133.1
(85)
263.4
(5)
405.9
(10)
316.6
(55)
3,276.9
(100)
2,850.2
(100)
200.3
(10)
8.0
(5)
176.2
(40)
991.8
(25)
Hay
963.6
(20)
49.4
(10)
95.5
(<5)
365.0
(55)
1,604.7
(65)
101.6
(10)
470.7
(30)
147.8
(5)
1,604.7
(10)
1,209.6
(40)
2,722.9
(90)
780.9
(65)
244.4
(40)
44.0
(<5)
1,978.2
(40)
268.2
(5)
Wheat Tobacco Potatoes
25.3
(40)
.2
(<5)
15.0
(10)
1.8
(40)
140.0
(85)
34.2
(70)
96.7
(70)
45.9
(5)
110.3
(20)
218.3
(55)
967.6
(95)
232.2
(100)
5.0
(50)
1.4
(<5)
37.4
(<5)
27.7
(5)
0
(0)
0
(0)
0
(0)
.3
(0)
142.0
(70)
0
(0)
19.3
(35)
0
(0)
1.9
(30)
0
(0)
19.3
(100)
0
(0)
0
(0)
0
(0)
0
(0)
0
(0)
110.5
(20)
0
(0)
45.5
(20)
2.0
(40)
0
(0)
0
(0)
0
(0)
0
(0)
0
(0)
57.3
(15)
0
(90)
0
(0)
0
(0)
0
(0)
0
(0)
0
(0)
a To the nearest
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8-17
Table 8-7 provides a ranking on the basis of production value of all
crops included in the inventory. From this table it is clear that
grain, corn and hay are most important, accounting for just over 50%
of total production value in eastern Canada.
A breakdown of the value and yield for each of these twelve crops by
deposition zone are provided in Tables 8-8a and 8-8b, respectively.
For many of the crops, more than 50% of their total yield is grown in
areas of high sulphate deposition, (over 40 kg/ha.yr). By contrast,
very small proportions (4% or less) are grown in areas experiencing
only 10 - 20 kg/ha.yr of sulphate deposition. It is evident that a
very significant proportion of Canada's agricultural crops are grown
in areas experiencing high deposition levels.
In order to obtain a better understanding of the geographical
distribution of these crops, Table 8-9 was prepared. This provides a
breakdown of production values for each province receiving more than
40 kg/ha.yr sulphate deposition. Appendix Tables 8-12 through 8-14
provide more detail.
Only Ontario and Quebec, which are significant agricultural
producers of all crops, have areas exposed to sulphate deposition in
excess of 40 kg/ha.yr. The most important crops in these areas
(based on value of production), are grain corn, hay and soybeans. In
the case of soybeans which is a small crop by volume, 95% of its
total volume of production is in the high deposition zone. This is
the highest proportion for any single crop, and all of this produc-
tion takes place in southwestern Ontario.
Overall, the most important crops in terms of value which are grown
in areas receiving 20 - 40 kg/ha.yr sulphate are hay, grain corn,
potatoes and tobacco. On a provincial basis, hay and grain corn are
a greater proportion of total value of production in Ontario and
Quebec, where potatoes are an important crop in the Maritime
provinces.
It is evident even from this preliminary analysis that a very large
proportion of eastern Canada's agricultural yields are grown in areas
of high and moderate deposition. In turn, the geographic distribu-
tion of crops varies so that certain crops represent a more signifi-
cant proportion of total value of production in each province. This
inventory has provided a preliminary overview of the agricultural
resources at risk, particularly in the high and medium deposition
zones. Better data on the responses of individual crop species to
these deposition regimes will provide the basis for a more accurate
quantification of the extent of risk.
-------
8-18
TABLE 8-7. RANKING OF CROPS IN EASTERN CANADA BY 1980 VALUE OF
PRODUCTION (1980 $ Cdn)
CROP
Grain Corn
Hay
Tobacco
Potatoes
Soybeans
Fodder Corn
Wheat
Oats
Barley
Cabbage
Lettuce
Spinach
TOTAL
1980 $ Value
(103)
794,173
658,550
333,821
279,940
228,499
253,084
128,605
100,595
88,756
15,878
9,332
1,222
2,892,495
Percent
of Total
28
24
12
10
8
6
5
4
3
<1
<1
<1
Cumulative
Percent
28
52
64
74
82
88
92
96
99
100
100
100
Production
(Metric Tons)
5,990.0
13,278.0
115.0
1,843.6
962.9
12,984.4
1,207.2
796.9
714.6
119.6
35,793.0
3,003.0
Source: Appendix Tables 8-12 and 8-13
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8-19
TABLE 8-8a. VALUE AND PERCENTAGE OF TOTAL 1980 YIELD OF EACH CROP
IN EASTERN CANADA BY DEPOSITION REGIME
Value 1980 $
(Percent of 1980
CROP
Grain Corn
Hay
Tobacco
Potatoes
Soybeans
Fodder Corn
Wheat
Oats
Barley
Cabbage
Lettuce
Spinach
Sulphate
Deposition
103
Yield)
kg/ ha
<10 10-20 20-40
0
(0)
0 21,
(0)
0
(0)
0 1,
(0)
0
(0)
0
(0)
0
(0)
0 3,
(0)
0 3,
(0)
0
(0)
0
(0)
0
(0)
0 226
(0)
271 354
(4)
0 159
(0)
607 217
0 10
(0)
25 104
(40
567,191
(62)
283,184
(43)
174,329
(51)
61,123
(21)
218,070
(95)
148,449
(56)
88,020
(67)
34,372
(31)
49,538
(52)
2,867
(20)
2,981
(25)
374
(34)
Total
794,173
658,550
333,821
279,940
228,499
253,084
128,605
100,595
88,756
15,878
9,332
1,222
-------
TABLE 8-8b.
1980 YIELD
DEPOSITION
OF MAJOR CROPS
REGIME
IN EASTERN
8-20
CANADA BY
Metric Tons 103
(Percent of
1980 Yield)
Sulphate Deposition kg/ha.
CROP
Grain Corn
Hay
Tobacco
Potatoes
Soybeans
Fodder Corn
Wheat
Oats
Barley
Cabbage
Lettuce
Spinach
<10
0
(0)
0
(0)
0
(0)
0
(0)
0
(0)
0
(0)
0
(0)
0
(0)
0
(0)
0
(0)
0
(0)
0
(0)
10-20
0
(0)
419.0
(3)
0
(0)
7.6
0
(0)
1.4
<
24.9
(3)
24.6
(4)
.2
(<1)
0
(0)
0
(0)
20-40
2,274.0
(38)
7,087.0
(53)
57.0
(50)
1,452.0
(79)
43.9
(5)
5,667.0
(44)
394.0
(33)
524.0
(66)
316.0
(44)
96.0
(80)
26.7
(75)
2.0
(66)
yr
> 40 Total
3,716.0 5,990.0
(62)
5,772.0 13,278.0
(44)
58.0 115.0
(50)
384.0 1,843.6
(21)
919.0 962.9
(95)
7,316.0 12,984.4
(56)
809.0 1,207.2
(67)
248.0 796.9
(31)
374.0 714.6
(52)
23.4 119.6
(20)
9.1 35.8
(25)
1.0 3.0
(34)
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8-21
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8-22
8.4.1 U.S. Forest Resources
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8.4 FOREST RESOURCES
Data on forest resources are aggregated and information on individual I
tree species are not provided in any terms. For the U.S., the
forests are differentiated into two categories (i.e., hardwood and
softwood), while in Canada there are three categories (i.e., hard- •
wood, mixed and softwood). Quantitative information of the total •
volume of forest resources (yield), its growth (or annual yield) and
value is provided in the inventory for each state or province and •
deposition regime. |g
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As of 1977, total commercial forest land in the U.S. was 197 x 106
hectares, and the total timber volume was 22.6 x 10^ m3 (USDA B
1978b). The annual growth was 350 x 10*> m3 of softwoods and 270 m
x ID** m3 of hardwoods. The average stumpage price in 1978
dollars was $27.50 per m3 for softwoods and $8.60 per m3 for •
hardwoods (Ulrich 1981). Combining the annual growth and appropriate g
value estimates gives a value of $11.9 billion to the net annual
growth. _
The total forest land area in those states east of the 100° meridian '
was 145 x 10" hectares, and the total timber volume was 9.8 x 10^
m3. The annual growth was 224 x 10^ m3 of softwood and 252 x B
10^ m3 of hardwoods. Combining the annual growth and appropriate |
value estimates results in a value of $8.3 billion for the net annual
growth in the eastern United States. •
The total forest volume and annual growth grouped by the two higher
deposition categories are displayed in the Appendix to this section. —
Note that the data on annual growth are incomplete for several •
states. These data are not available on a county basis, so they did •
not appear in the data summary.
The volume and growth increments show a similar distribution among |
the four deposition categories (Table 8-10). Approximately 10% of
the hardwood and softwood growth is found in areas of highest deposi- «
tions and over 75% of the hardwood and softwood growth is found in •
areas of moderate deposition.
Although the aggregate 38 state data show that only 15% of the forest I
volume is exposed to sulphate deposition greater than 40 kg/ha. yr, •
individual state data show a different picture (Table 8-11). A
significant portion of the forest areas in the states of Arkansas, •
Ohio and Texas receives sulphate deposition greater than 40 kg/ha. yr. f
In nine states 30% or more of the forest area receives sulphate
deposition greater than 40 kg/ha. yr. _
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8-23
TABLE 8-10.
U.S. HARDWOOD AND SOFTWOOD VOLUME AND GROWTH
(Olson et al. 1980)
Volume
Hardwood Volume3
Softwood Volume3
Total Volume
Growth
Hardwood Growtha
Softwood Growtha
Total Growth
<10
0
(0)
40
(0)
40
(0)
(0)
(0)
(0)
(Percent
Sulphate
10-20
780
(10)
480
(15)
1,290
(10)
10
(10)
20
(15)
30
(15)
m3 106
of 38 State
Deposition
20-40
4,040
(70)
2,690
(75)
6,930 1
(70)
110
(80)
140
(75)
250
(75)
Total )b
kg/ha. yr
>40
940
(15)
480
(10)
,540
(15)
20
(10)
20
(10)
40
(10)
Total
5,750
3,690
9,800
140
190
330
a These numbers may not add to total numbers because data for some
states do not distinguish between hardwood and softwood.
To the nearest 5%.
-------
8-24
TABLE 8-11. U.S. FOREST VOLUME BY STATE RECEIVING GREATER THAN
40 kg/ha.yr SULPHATE DEPOSITION (Olson et al. 1980)
Volume m
REGION II
New York
REGION III
Maryland
Pennsylvania
West Virginia
REGION IV
Kentucky
Mississippi
Tennessee
REGION V
Illinois
Indiana
Michigan
Ohio
REGION VI
Arkansas
Louisiana
Oklahoma
Texas
REGION VII
Missouri
3 . 10^ (Percent of
41,900
7,800
180,200
125,400
123,300
45,400
102,000
14,200
5,300
25,800
113,300
195,300
194,200
18,500
308,900
35,000
State Total)3
(10)
(10)
(30)
(35)
(40)
(10)
(35)
(20)
(5)
(5)
(95)
(80)
(40)
(30)
(100)
(20)
See Appendix Table 8-11 for growth data.
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8-25
8.4.2 Canadian Forest Resources
The wood industry is an important component of the Canadian economy.
Forest industries valued at $22 billion Cdn. annually constitute
Canada's largest manufacturing industry as well as the largest single
contributor to the positive side of our balance of payments (Sidor
1981). One in ten Canadian jobs depends on the forestry sector.
The importance of the forest industry in the eastern Canadian
provinces to the wood industry is substantial. Approximately 35% of
the country's total productive forest land lies within the boundaries
of eastern Canada. Further, the eastern Canadian provinces accounted
for about 64% ($3.5 billion Cdn.) of Canada's total value added in
the forest industry (Sidor 1981). Total value of the annual forest
growth of 150,241,000 m3 is estimated to be $3.9 billion Cdn. This
is based on an average wood value of $26.25 Cdn. per m3 (1981).
Figure 8-3 illustrates the pattern of acidic deposition (kg SO^"
/ha.yr) and forest type. In eastern Canada higher levels of
8042- are most often associated with deciduous forests and
lower 8042- levels with coniferous.
Table 8-12 lists the annual growth by forest type and deposition
regime. Although only 4% (5,436 m^.lO3) of the annual growth
occurs in areas receiving more than 40 kg/ha.yr sulphate deposition
this does represent 10% of the hardwood annual yield. Although
deposition exceeding 40 kg/ha.yr sulphate affects the smallest area
of forested land (2,048.000 ha), this is the area of highest mean
annual increment (2.1 m-Vha) affecting mixed and hardwood forests.
The bulk of hardwood and mixed wood growth occurs in areas receiving
20 - 40 kg/ha.yr sulphate, representing 64% and 70% of annual growth
by forest type, respectively.
The provincial summary (Table 8-13) illustrates the geographic
variation in annual growth. While only 41% and 30% of the annual
growth for Quebec and Ontario are receiving more than 20 kg/ha.yr of
sulphate, 100% of the annual growth in the Maritime provinces are
under the moderate deposition regime. Sixty-seven percent of
Newfoundland's forest growth occurs under similar conditions
receiving 20 - 40 kg/ha.yr of sulphate.
8.5 MAN-MADE STRUCTURES
Man-made materials can be grouped into three classes (i.e., metals,
masonary and organic materials). Organic materials include paints,
coatings, textiles and wood. Materials within each one of these
classes behaves differently when exposed to air and water
pollutants.
-------
8-26
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8-27
TABLE 8-12. HARDWOOD, SOFTWOOD AND MIXED WOOD ANNUAL GROWTH
(m3.!03) IN EASTERN CANADA BY DEPOSITION REGIME
Annual Growth m
3.103
(Percent of Total)
Sulphate Deposition
10-20 20-40
Hardwooda
Sof twooda
Mixed Wooda
Total Annual
Growth
5,082
(26)
66,753
(71)
8,327
(23)
80,162
(53)
12,491
(64)
26,814
(28)
25,341
(70)
64,646
(43)
kg/ha. yr
>40
1,825
(10)
815
(1)
2,760
(8)
5,400
(4)
Tot alb
19,398
94,382
36,428
150,208
a Hardwood and softwood forests contain 70% or more of the
specified type. Mixed wood forests contain less than 70% of either
hardwood or softwood species.
b Total annual growth does not include data on regions receiving
less than 10 kg/ha.yr sulphate desposition.
-------
8-28
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TABLE 8-13.
ANNUAL FOREST GROWTH (m3.103) BY PROVINCE RECEIVING
GREATER THAN 20 kg/ha.yr SULPHATE DEPOSITION
Ontario
Quebec
New Brunswick
Nova Scotia
Prince Edward Island
Newfoundland
Annual Growth mj
(Percent of Provincial Total)
Sulphate Deposition kg/ha.yr
20-40 >40
12,544(28)
29,659(36)
11,474(100)
7,077(100)
443(100)
3,456(67)
998(2)
4,438(5)
0(0)
0(0)
0(0)
0(0)
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8-29
An inventory of man-made structures must distinguish between renew-
able materials and cultural materials. Renewable materials are those
which are easily replaceable. They include such items as surface
coatings (paint), chain link fence and galvanized roofing. Cultural
materials are those which are difficult to replace because of the
scarcity of material and requirements for skilled craftsmen to
recreate the resource. They include such items as sculptured stone
and metals and dimension stone. There are several materials (e.g.,
adobe, plaster concrete and unit masonry) which could fall into
either category depending on the craftsmanship requirements.
There is no national inventory of renewable materials which are
susceptible to acidic deposition. Past efforts to create a national
inventory for urban areas have combined per capita material
estimates, based on limited survey data, and census data on
population distribution. Although this approach to creating an
inventory of renewable resources has been used, the resulting urban
inventories are of questionable value due to the uncertainties
associated with the per capita material estimates (Koontz et al.
1981; Stankunas et al. 1981). Results in two Standard Metropolitan
Statistical Areas (SMSA) indicate that the area of urban development
and local availability of materials are important factors in the
distribution of material quantities. There are no estimates of
renewable materials in rural areas. Until additional survey work is
complete, the Work Group cannot provide an acceptable national
inventory of renewable materials or an estimate of materials by
sulphur deposition regimes.
Complete national inventories of cultural materials are not available
for either Canada or the U.S. Such national inventories would
include all significant cultural materials, both historic and
contemporary. The only inventory of cultural materials that can be
assembled at this time is one of major historical resources. Both
the U.S. and Canada maintain lists of significant historic sites and
artifacts. The limitations of these sources are that they are
incomplete in not listing all significant materials, such as sculpted
stone and metals in urban areas and that the data on those items is
not always adequate for an analysis of potential damage from air
pollution.
8.5.1 U.S. Historic Inventory
The Work Group compiled a general U.S. inventory of historic
resources based on Federal data sources. These sources were the
National Register of Historic Places (U.S. Federal Register 1979,
1980, 1981, 1982), National Historic Landmarks (USDI 1976) and
National Historic Parks (USDI 1982). The National Register of
Historic Places includes sites because of their association with an
event or person, of their architectural or engineering qualities, or
of their potential contribution to historic studies. It consists of
approximately 26,000 sites and is the most comprehensive of all three
-------
8-30
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sources. National Historic Landmarks are those historic places
designated by the Secretary of Interior to be of national signifi-
cance. It consists of about 1,500 sites. National Historic Parks is •
used as a generic designation of properties of national sigificance •
owned by the Federal government. It includes historic sites,
military battlefields and historic monuments and numbers about 150
sites.
I
The Work Group cross-classified historical resources by three ambient •
S0£ categories at the county level. The three ambient categories •
at the county level were chosen arbitrarily because there is no S02
standard designated for protecting materials resources nor did the
Material Effects Work Group establish damage functions for material •
damages (see Section 5). The results of the tabulation are •
summarized by state in Table 8-14. The sum of all such sites in the
three ambient categories is the total for each state. No attempt was •
made to distinguish a major and minor site within each category so •
the numbers should be interpreted with great care.
Only historic sites in seven out of the 38 states under consideration •
experience ambient S02 concentrations greater than 80 yg/m3. ™
Within those states, the majority of historic places, landmarks and
monuments are located in counties with ambient S02 concentrations •
less than 60 yg/m3. Only in the states of Illinois and New York •
are there more than 20% of the historic sites in counties with
ambient concentrations greater than 80 yg/m3. in total, approxi- •
mately 3% of the historic places, 3% of the historic landmarks and 2% •
of the historic parks and monuments are located in counties with
ambient concentrations greater than 80 yg/m3.
8.5.2 Canadian Historic Inventory
For the purposes of this inventory, there are three categories (i.e., I
national historic sites, buildings and museums, and monuments and
parks). Data for all of these categories are available only for _
Ontario. For the other provinces of eastern Canada, only national •
historic sites are included. These have been further subdivided into *
two deposition regimes; under 40 and over 40 kg/ha.yr. Although with
structures, concentrations of S02 (in yg/m3) is perhaps more I
appropriate, no ambient air quality data are available from which •
areas of uniform concentrations can be drawn. Generally, higher
levels (i.e., above 55 yg/m3 of 802) are found in the major •
cities, and even then the annual averages are much lower. •
The inventory data presented in Table 8-15 indicate that the majority
of national historic sites in areas of high deposition are found in •
Ontario, with the balance in Quebec, in the vicinity of Quebec City, •
the oldest settlement in eastern North America.
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8-31
TABLE 8-14.
U.S. HISTORIC SITES BY AMBIENT S02 (yg/m3) (USDI
1976; USEPA 1980)
REGION I
Connecticut
Maine
Massachusetts
New Hampshire
Rhode Island
Vermont
REGION II
New Jersey
New York
REGION III
Delaware
Dist. of Col.
Maryland
Pennsylvania 1
Virginia
W. Virginia
REGION IV
Alabama
Florida
Georgia
Kentucky
Mississippi
North Carolina
South Carolina
Tennessee
REGION V
Illinois
Indiana
Michigan
Ohio 1
Wisconsin
REGION VI
Arkansas
Louisiana
Oklahoma
Texas
Historic
Places
<60 60-80 >80
459
384 170 36
649
226 - 6
324
297
485
827 126 211
273
224
510
,201 - 163
916
255 - 2
303
363
705
543 142
426
617
460
465 1
419 - 148
326 27 9
468
,193 404 16
603
409
367
357
731
Historic
Landmarks
<60 60-80 >80
29
23 - 1
105
16 - 12
23
10
23
87 - 16
11
37
42
66
78
3 - -
14
15
25
11 3 -
15
19
59
18
18 - 12
7 - 1
7 - -
25 - 1
12
0 — —
34
13
21
Historic
<60
-
-
9
1
2
-
4
9
-
12
9
12
12
2
2
-
6
3
3
5
3
6
1
1
-
1
—
3
3
1
5
Parks
60-80
-
-
-
-
-
-
-
1
-
-
-
5
-
—
-
-
-
-
-
-
-
-
-
-
-
-
—
-
-
-
>80
-
-
-
-
-
-
-
3
-
-
-
-
-
—
-
-
-
-
-
-
-
-
-
-
-
-
—
-
-
-
-------
8-32
TABLE 8-14. CONTINUED
REGION VII
Iowa
Kansas
Missouri
Nebraska
REGION VIII
North Dakota
South Dakota
TOTAL
<60
497
292
550
279
106
209
18,766
Historic
Places
60-80 >80
22
-
20
- -
-
-
912 591
Historic
Landmarks
<60 60-80 >80
9
14
19 2 -
14
1 - -
10
948 5 33
Historic
Parks
<60 60-80
1
2
3
1
2
-
124 6
>80
-
-
-
-
-
-
3
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8-33
TABLE 8-15. CANADIAN HISTORIC INVENTORY BY PROVINCE AND DEPOSITION
— (OMCR 1978; Parks Canada 1981)
• National Buildings Monuments
™ Sulphate Historic Sites and Museums and Parks
Deposition <40 >40 <40 >40 <40 >40
fl Province (kg/ha.yr)
Ontario 21 23 129 72 8 3
J Quebec 13 5 N/A N/A
_ Prince Edward
• Island 4 N/A N/A
Nova Scotia 13 N/A N/A
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New Brunswick 2 N/A N/A
Newfoundland 6 N/A N/A
TOTAL 59 28
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8-34
Quebec Region, Ste-Foy, P.Q. (unpublished)
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8.6 REFERENCES
Bickerstaff, A.; Wallace, W.L.; and Evert, F. 1981. Growth of •
forests in Canada, Part 2. A quantitative description of the •
land base and the mean annual increment. Information Report
Pl-X-1, Canadian Forestry Service, Environment Canada, Ottawa,
Ont. I
Cox, E.T. 1978. Counts and measurements of Ontario lakes, 1978.
Ontario Ministry of Natural Resources, Toronto, Ont. 114 pp. •
Environment Canada. 198la. Ecodistrict maps and descriptions for
the Atlantic provinces. Lands Directorate, Atlantic Region, «
Halifax, N.S. (unpublished) •
. 1981b. Les ecodistricts du Quebec. Lands Directorate,
I
Koontz, M.D.; McFadden, J.E.; and Haynie, F.H. 1981. Estimation of
total surface and spatial distribution of exposed building •
materials from commonly available information for U.S. |
metropolitan areas. In Proc. Second Int. Conf. on the
Durability of Building Materials and Components, 1981. _
Lucas, A.E., and Cowell, D.W. 1982. A regional assessment of
sensitivity to acidic deposition for eastern Canada. Presented
at Symp. Acid Precipitation. Am. Chem. Soc., Las Vegas, NV., I
1982. •
New Brunswick Department of Agriculture and Rural Development. •
Telephone Inquiry, March 3, 1982. I
Newfoundland Department of Agriculture and Forestry. Telephone _
Inquiry, March 3, 1982. •
Nova Scotia Department of Agriculture and Marketing. Telephone
Inquiry, March 3, 1982. •
Olson, R.J.; Johnson, D.W.; and Shriner, D.S. 1982. Regional
assessment of potential sensitivity of soils in the eastern •
United States to acid precipitation. Oak Ridge National I
Laboratory, Oak Ridge, TN. 31 pp. (in press)
Ontario Ministry of Agriculture and Food (OMAF). 1981. Agricultural •
Statistics for Ontario, 1980. Publication 20, Toronto, Ont. ™
Ontario Ministry of Agriculture and Food (OMAF). 1981. Seasonal •
fruit and vegetable report, fruit and vegetable production in |
Ontario, 1980. Toronto, Ont.
Ontario Ministry of Culture and Recreation (OMCR). 1978. Ontario I
historic sites, museums, galleries and plaques. Toronto, Ont.
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8-35
Parks Canada. 1981. List of all positive recommendations of the
historic sites and monuments board of Canada, from 1919 to
June 1, 1981. Ottawa, Ont. (unpublished)
Prince Edward Island, Department of Agriculture and Forestry.
Telephone Inquiry, March 3, 1982.
Quebec Bureau de la Statistique (QBS). 1978. Grandes cultures,
1977. Superficie, production et valeur. Quebec, P.Q.
Sidor, N. 1981. Forest industry development policies: industrial
strategy or corporate welfare? Canadian Centre for Policy
Alternatives, Pub. #3, Inaugural Conference Proceedings.
Stankounas, A.R.; Unites, D.F.; and McCarthy, E.F. 1981. Air
pollution damage to man-made materials: physical and economic
estimates^Final Report to EPRI #RT1004.
Statistics Canada. 1976. Census of agriculture. Census of Canada,
Volume XI, Ottawa, Ont.:
1976a. Quebec. Catalogue 96-805.
1976b. Ontario. Catalogue 96-806.
1976c. Newfoundland. Catalogue 96-801.
1976d. Prince Edward Island. Catalogue 96-802.
1976e. Nova Scotia. Catalogue 96-803.
1976f. New Brunswick. Catalogue 96-804.
. 1981a. Field crop reporting. Catalogue 22-002, Ottawa,
Ont.
1981b. Fruit and vegetable production. Catalogue
22-003, Ottawa, Ont.
Ulrich, A. 1981. U.S. timber production, trade, consumption and
price statistics, 1950-80. USDA Forest Service, Washington,
DC.
U.S. Department of Agriculture (USDA). 1971. Basic statistics -
national inventory of soil and water conservation needs, 1967.
Statistics Bulletin No. 461, USDA, Washington, DC. 211 pp.
1978a. National resource inventories. USDA Soil
Conservation Service, Washington, DC.
. 1978b. Forest Statistics of the U.S. USDA Forest
Service, Washington, DC.
. 1979. Volume and value of sawtimber stumpage sold from
national forests by selected species and region, 1978. USDA
Forest Service, Washington, DC. (mimeo)
. 1980. Agricultural Statistics. Washington, DC.
-------
8-36
U.S. Federal Register. Nation register of historic places.
1979. February 6, 1979.
1980. March 18, 1980.
1981. February 3, 1981.
1982. February 2, 1982.
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I
U.S. Department of Commerce (USDC). 1979. Census of agriculture,
1978. Preliminary file, Technical documentation, Bureau of _
Census, USDC, Washington, DC. I
U.S. Department of Interior (USDI). 1976. National Historic
Landmarks, Washington, DC. •
. 1982. List of classified structures. National Park
Service. Computer Printout. •
U.S. Environmental Protection Agency (USEPA). 1980. Ambient S02
data 1979-1980. Office of Air Quality, Planning and Standards.
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APPENDIX TABLE 8-1.
8-37
U.S. AQUATIC RESOURCES BY STATE AND SENSITIVITY
CATEGORY (km2) 10-20 kg/ha.yr SULPHATE
DEPOSITION (USDA 1971, 1978)
REGION III
Virginia
West Virginia
REGION IV
Florida
Georgia
REGION V
Illinois*
Michigan
Minnesota
Wisconsin
REGION VI
Oklahoma
Texas
REGION VII
Iowa a
Kansas
Missouri^
Nebraska*
REGION VIII
North Dakota
South Dakota
TOTAL
Low
Potential
to Reduce
Acidity
10
10
0
280
0
1,350
1,460
2,310
0
0
0
0
0
0
0
0
5,420
Moderate
Potential
to Reduce
Acidity
20
0
100
210
0
0
8,440
1,240
0
0
0
0
0
0
620
410
11,040
High
Potential
to Reduce
Acidity
40
0
9,560
180
0
190
890
240
1,170
1,060
0
1,000
0
0
200
1,210
15,740
Total
Surface
Water Area
in the State
2,950
480
13,130
3,000
2,110
4,180
13,810
5,420
4,440
15 -,260
990
3,140
2,770
2,330
4,540
4,240
82,790
a These states are included, even though they do not have surface
water area in counties falling into one of the three sensitivity
categories, because they have surface water area falling into the
urban/agricultural category receiving 10-20 kg sulphate/ha.yr.
-------
APPENDIX TABLE 8-2.
8-38
U.S. AQUATIC RESOURCES BY STATE AND SENSITIVITY
CATEGORY (km2) 20-40 kg/ha.yr SULPHATE
DEPOSITION (USDA 1971, 1978)
REGION I
Connecticut
Maine
Massachusetts
New Hampshire
Rhode Island
Vermont
REGION II
New Jersey
New York
REGION III
Delaware
Dist. of Columbiab
Maryland
Pennsylvania
Virginia
W. Virginia
REGION IV
Alabama
Florida
Georgia
Kentucky
Mississippi
North Carolina
South Carolina
REGION V
Illinois
Indiana
Michigan
Ohio
Wisconsin^
Tennessee
REGION VI
Arkansas
Louisiana
Oklahoma
Texas
Low
Potential
to Reduce
Acidity
340
6,010
970
860
340
720
150
1,930
0
0
0
40
70
110
10
0
250
30
60
310
0
0
0
1,650
0
0
90
0
0
0
0
Moderate
Potential
to Reduce
Acidity
0
0
0
0
0
0
660
1,050
0
0
1,020
300
1,600
20
1,280
410
1,660
140
970
9,010
640
0
40
0
0
0
100
180
0
880
10
High
Potential
to Reduce
Acidity
0
0
0
0
0
0
10
520
0
0
80
400
840
120
1,590
960
560
410
620
1,110
2,100
90
150
140
20
0
1,440
500
10,050
1,330
3,040
Total
Surface
Water Area
in the State
500
6,010
1,100
860
460
1,000
920
5,260
330
0
1,900
1,620
2,950
480
3,040
13,130
3,000
2,320
2,470
10,480
2,750
2,110
930
4,180
970
5,420
3,140
3,670
11,390
4,440
15,260
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8-39
APPENDIX TABLE 8-2. CONTINUED
REGION VII
lowab
Kansas^
Missouri
Low
Potential
to Reduce
Acidity
0
0
0
Moderate
Potential
to Reduce
Acidity
0
0
0
High
Potential
to Reduce
Acidity
0
0
1,320
Total
Surface
Water Area
in the State
990
3,140
2,770
TOTAL 13,940 19,970 27,400 107,890
a The states of Minnesota (13,810 km2), Nebraska (2,330 km2),
North Dakota (4,540 km2) and South Dakota (4,240 km2) received
less than 20 kg/ha.yr sulphate deposition in 1979-80.
" These states are included, even though they do not have surface
water area in counties falling into one of the three sensitivity
categories because they have surface water area falling into the
urban/ agricultural category receiving 20-40 kg sulphate/ha.yr.
-------
APPENDIX TABLE 8-3.
8-40
U.S. AQUATIC RESOURCES BY STATE AND ACID
SENSITIVITY CATEGORY (km2) FOR GREATER THAN 40
kg/ha.yr SULPHATE DEPOSITION (USDA 1971, 1978)
REGION II
New York
REGION III
Maryland
Pennsylvania
W. Virginia
REGION IV
Kentucky
Mississippi
Tennessee
REGION V
Illinois
Indiana
Michigan
Ohio
REGION VI
Arkansas
Louisiana
Oklahoma
Texas
REGION VII
Missouri
TOTAL
Low
Potential
to Reduce
Acidity
0
20
0
10
0
0
0
0
10
40
0
130
0
0
0
0
210
Moderate
Potential
to Reduce
Acidity
200
0
250
0
260
90
10
0
0
0
0
1,170
0
210
0
0
2,190
High
Potential
to Reduce
Acidity
0
0
220
210
480
10
640
80
0
40
360
400
1,050
0
4,050
80
7,620
Total
Surface
Water Area
in the State
5,260
1,900
1,620
480
2,320
2,470
3,140
2,110
930
4,180
970
3,670
11,390
4,440
15,260
2,770
62,910
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8-41
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I
I
I
I
I
I
I
I
I
I
I
I
I
I
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-------
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I
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I
APPENDIX TABLE 8-8.
8-47
U.S. FOREST RESOURCES IN AREAS RECEIVING 20-40
kg/ha.yr SULPHATE DEPOSITION - VOLUME (m3.103)
(Olson et al. 1980)
REGION I
Connecticut
Maine
Massachusetts
New Hampshire
Rhode Island
Vermont
REGION II
New Jersey
New York
REGION III
Delaware
District of
Columbia
Maryland
Pennsylvania
Virginia
W. Virginia
REGION IV
Alabama
Florida
Georgia
Kentucky
Mississippi
North Carolina
South Carolina
Tennessee
REGION V
Illinois
Indiana
Michigan
Ohio
Wisconsin
Total
Volume
66,600
601,800
96,100
186,300
9,800
133,900
41,600
303,600
16,600
-
91,100
445,600
509,600
251,800
572,300
75,900
647,000
200,700
440,600
702,400
486,500
192,400
51,200
93,900
184,800
5,200
8,200
Softwood
Volume
10,100
418,000
35,900
88,900
2,100
48,600
7,300
84,100
5,200
-
22,000
28,500
144,000
17,300
319,500
48,100
370,300
-
239,700
294,000
252,800
44,300
100
1,700
33,400
100
500
Hardwood
Volume
56,500
183,800
60,200
97,400
7,700
85,300
34,300
219,500
11,400
-
69,000
417,200
365,600
234,500
252,800
27,800
276,700
-
201,000
408,400
233,700
148,100
51,100
92,200
151,500
5,100
7,700
Total Volume
in State3
66,600
601,800
96,100
186,300
9,800
133,900
41,600
345,400
16,600
-
98,900
625,800
548,600
382,700
572,300
367,100
714,600
324,000
486,000
702,400
486,500
294,400
66,400
99,200
425,400
118,500
315,400
-------
8-48
APPENDIX TABLE 8-8. CONTINUED
REGION VI
Arkansas
Louisiana
Oklahoma
Texas
REGION VII
Iowa
Kansas
Missouri
TOTAL
Total
Volume
56,700
278,600
39,900
6,900
3,500
600
127,600
6,929,500
Softwood
Volume
11,200
132,500
17,300
3,800
0
6,600
2,688,000
Hardwood
Volume
45,500
146,100
22,600
3,100
600
121,000
4,037,400
Total Volume
in State3
252,000
472,800
58,400
315,800
29,900
8,400
172,100
9,435,700
I
I
I
I
I
I
I
I
a Under all deposition regimes.
I
I
I
I
I
I
I
I
I
I
-------
8-49
APPENDIX TABLE 8-9. U.S. FOREST RESOURCES IN AREAS RECEIVING 20-40
kg/ha.yr SULPHATE DEPOSITION - GROWTH (m3.103)
(Olson et al. 1980)
Total
Growth
Softwood
Growth
Hardwood
Growth
REGION I
Connecticut 1,700
Maine 18,000
Massachusetts 3,100
New Hampshire 6,700
Rhode Island 300
Vermont 3,000
REGION II
New Jersey 700
New York
REGION III
Delaware 500
District of
Columbia
Maryland
Pennsylvania
Virginia 21,400
W. Virginia
300
14,000
1,200
3,600
100
1,300
200
1,300
6,500
1,400
4,000
2,000
3,140
200
1,700
500
400
14,900
Total Growth
in State3
1,700
18,000
3,200
6,700
300
3,000
700
500
23,000
REGION IV
Alabama
Florida
Georgia
Kentucky
Mississippi
North Carolina
South Carolina
Tennessee
REGION V
Illinois
Indiana
Michigan
Ohio
Wisconsin
33,600
4,400
40,800
-
26,100
31,900
27,300
9,400
1,900
-
-
-
22,300
3,400
29,400
-
15,800
15,200
17,500
2,500
0
-
-
-
11,300
1,000
11,400
-
10,300
16,700
9,800
6,900
1,900
-
-
-
33,600
21,500
44,500
-
28,700
31,900
27,300
14,400
2,400
-
-
-
-------
8-50
APPENDIX TABLE 8-9. CONTINUED
REGION VI
Arkansas
Louisiana
Oklahoma
Texas
REGION VII
Iowa
Kansas
Missouri
TOTAL
Total
Growth
-
14,300
2,400
300
-
-
2,700
251,500
Softwood
Growth
-
8,400
1,000
200
-
-
300
143,100
Hardwood
Growth
-
5,900
1,400
100
-
-
3,400
108,400
Total Growth
in State3
-
26,300
3,400
18,300
-
-
5,100
314,500
I
I
I
I
I
I
I
I
a Under all deposition regimes.
I
I
I
I
I
I
I
I
I
I
-------
I
I
I
I
I
I
I
I
I
I
I
I
I
I
I
I
I
I
I
8-51
APPENDIX TABLE 8-10.
U.S. FOREST RESOURCES IN AREAS RECEIVING
GREATER THAN 40 kg/ha.yr SULPHATE DEPOSITION
VOLUME (m3.103) (Olson et al. 1980)
REGION II
New York
REGION III
Maryland
Pennsylvania
W. Virginia
REGION IV
Kentucky
Mississippi
Tennessee
REGION V
Illinois
Indiana
Michigan
Ohio
REGION VI
Arkansas
Louisiana
Oklahoma
Texas
REGION VII
Missouri
TOTAL
Total
Volume
41,900
7,800
180,200
125,400
123,300
45,400
102,000
14,200
5,300
25,800
113,300
195,300
194,200
18,500
308,900
35,000
1,536,500
Softwood
Volume
6,000
400
8,600
10,500
13,200
6,600
400
0
1,000
3,200
80,900
123,500
11,300
205,700
3,900
475,300
Hardwood
Volume
35,900
7,400
171,600
114,900
32,200
95,300
13,800
5,300
24,800
110,100
144,400
70,700
7,200
103,200
31,100
937,800
Total Volume
in State3
345,400
98,900
625,800
382,700
324,000
486,000
294,400
66,400
99,200
425,400
118,500
252,000
472,800
58,400
315,800
172,100
45,378,000
Under all deposition regimes.
-------
8-52
APPENDIX TABLE 8-11.
U.S. FOREST RESOURCES IN AREAS RECEIVING
GREATER THAN 40 kg/ha.yr SULPHATE DEPOSITION
GROWTH (m3.103) (Olson et al. 1980)
REGION II
New York
REGION III
Maryland
Pennsylvania
W. Virginia
REGION IV
Kentucky
Mississippi
Tennessee
REGION V
Illinois
Indiana
Michigan
Ohio
REGION VI
Arkansas
Oklahoma
Louisiana
Texas
REGION VII
Missouri
TOTAL
Total
Growth
_
-
2,600
5,100
400
-
-
—
-
900
12,000
18,000
1,100
40,100
Softwood
Growth
-
-
900
500
0
-
-
—
-
500
8,700
12,000
100
22,800
Hardwood Total Growth
Growth in State3
-
-
1,700
4,600
400
-
-
—
-
400
3,300
6,000
1,000
17,300
-
-
28,700
14,400
2,400
-
-
—
-
3,400
26,300
18,300
5,100
98,600
a Under all deposition regimes.
I
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8-53
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SECTION 9
LIMING
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9-1
SECTION 9
LIMING
9.1 INTRODUCTION
This chapter reviews the potential of adding neutralizing or
buffering materials to correct or modify the adverse effects
associated with acidic deposition. Such activities are commonly
called "liming" and they are limited in practice to correcting to
some degree adverse effects on aquatic, terrestrial and drinking
water systems. It is not possible to use liming to mitigate effects
on materials, visibility, and adverse health effects resulting from
direct inhalation of airborne pollutants. Each section of this
chapter discusses the effectiveness of liming for the particular
system and then the unit cost associated with liming that system.
9.2 AQUATIC
It has been shown in many areas of the world that acid loadings due
to long range transport are capable of acidifying surface waters. In
theory however, even the most acidic loadings could periodically be
neutralized if limestone were added to the affected systems in
amounts ranging from 50 to 100 kg/ha.yr. In areas with calcareous
soils, this amount of neutralizing capacity is available inherently
for very long periods of time (i.e., 1 cm of soil covering 1 ha is
about 150 metric tons, which is capable of neutralizing present
maximum acid loading for about 3,000 years). In hard rock areas with
little or no calcareous soil some present acid loadings cannot be
neutralized fast enough resulting in acidic runoff. In order to
reverse or prevent the resulting effect, at least five different
jurisdictions (Sweden, Norway, New York State, Nova Scotia, and
Ontario) have added neutralizing agents to surface water systems.
The numbers of lakes and rivers treated and the methods used in the
application of neutralizing agents vary greatly from area to area.
Limestone is most often used although other chemicals have been
tried. The term "liming" is used to describe artificial
neutralization regardless of the chemical or chemicals actually
used.
9.2.1 Liming as a Mitigative Measure
In certain cases, a species or a unique race of organisms may be
threatened by acidification of its natural habitat. In these cases,
liming or other mitigative measures might be undertaken on lakes,
rivers or parts of rivers in order to preserve a population of the
endangered organism. Very small populations become inbred and so
the preserved habitat must be large enough to support a reasonably
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large population. The populations saved in this way might play an
important part in the restoration of fisheries to waters where fish
have been exterminated by acidification or other causes. «
The future value of any species or organism cannot be foreseen.
Therefore, the extinction of any species could be a great loss to
man. As the extinction of a species can never be remedied, the •
threat of extinction of any species by acid rain would justify lake •
liming or virtually any other feasible protective measure regardless
of immediate costs or benefits. •
If a population of fish is considered an especially suitable source
of stock for the rehabilitation of acidified rivers or lakes, then ^
liming or other measures to protect its natural habitat would be I
justified to protect it from acidification. Liming of an acidified
habitat would also be justified if it were inhabited by a population
which was genetically unique and consequently for which no replace- •
ment could be found if it were exterminated. •
9.2.2 Liming Programs I
9.2.2.1 Sweden
Sweden has conducted the greatest number of experiments on lake ™
liming of any country. A 5-year program designed to evaluate both
lake and stream liming was completed in 1981 and a final summary •
report was prepared (National Fisheries Board and National ||
Environmental Protection Board 1981). During the 5-year period, 304
projects were started which involved over 700 lakes and streams. M
While there were some negative aspects to the results, the program •
was deemed to be a success by the National Board of Fisheries.
Success was generally measured in terms of a favourable response in
the sport fish, mainly salmon, trout and arctic char, although some •
lakes were treated with environmental conservation as the prime |
objective. Limestone has been applied at rates of 100-200 kg/ha of
lake surface which corresponds to 50-75 kg/ha of watershed per year. •
Application on land required up to 100 times this amount to give an •
acceptable runoff quality. Application directly to water was found
to be the most economical treatment method (Bengtsson et al. 1980). _
Hultberg and Andersson (1982) reported detailed studies on six lakes 9
in two areas of Sweden. Four lakes were limed and two held as
reference lakes. Although they reported favourable biological B
results, there was concern over continued biological and chemical |
damage from liming, resulting from the input of aluminum in acidified
runoff from the watershed. Most of the lakes and streams had a •
relatively small number of species of fish (three or four). The •
improved water quality generally resulted in increased numbers of
fish, and hence an improved sport fishery. Bengtsson (pers. comm.)
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reported that, in some cases, nonsport fish species responded the
most dramatically.
Bengtsson et al. (1980) summarized some of the problems with lake
neutralization as follows:
"Obtaining yearly leaching of a certain amount of bicarbonate or
an acceptable yearly pH-medium value is not the problem. The
problem is to keep an acceptable value at high flow (i.e.,
during snowmelt). At this time the lake waters become highly
stratified with the cold acid melting-water on top. As a result
it does not mix with the water below which is of better
quality.
The running waters carrying this melting-water represent an even
greater problem. To neutralize the acidified melting-water,
either large overdosing is needed when applied to water or on
land or every year lime has to be applied on the snow pack.
Moreover, the acidification is not just a pH-problem but is also
coupled to the anthropogenic pollution of metals deposited from
the atmosphere and the increased leaching of metals from acidi-
fied soils.
The toxicity of most metals is higher in neutral than in acid
water. Thus, when liming an acid lake the organisms suffer a
transition period before the metals have precipitated. Aluminum
leached from the soil is highly aggressive to fish gills in the
pH range 4.5-6 and liming has even killed salmon and trout
when the aim was to save the fish."
The situation in North America may be even more complicated because
generally, the lakes contain more species of fish than many of the
Scandinavian lakes. The potential for disruption of the aquatic food
chain is greater. It could happen that fish populations would
survive in the treated lakes but the normal distribution of species
might be altered. For example, inputs of aluminum might disrupt the
life cycle of some species more than others, changing the ecological
balance among species.
9.2.2.2 Norway
The Norwegian government is conducting a liming project at Lake
Hovvatn in southern Norway. The lake is about one square kilometre
and has a mean residence time of 1.1 years. The drainage basin is
interspersed forest and bog with numerous granite outcroppings. The
project was begun in May 1980 with background sampling at two month
intervals at five lake stations and five inlet streams. Analyses
include pH, alkalinity, conductivity, all major ions and metals.
Zooplankton and phytoplankton are also being monitored. In March
1981, the lake was treated with 240 metric tons of agricultural
limestone spread on the ice near the shore. As the ice melted in
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State
Kansas
Massachusetts
New Jersey
New York
West Virginia
Wisconsin
Ponded Waters
Per Average
5
2-3
2
7
1
2
Treated
Year
(10 ha)
(81 ha)
( 8 ha)
(81 ha)
(17 ha)
(12 ha)
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April the pH went from 4.5 to 6.0. Sampling and analyses are
expected to continue for years. The lake was stocked with brown
trout in June 1981. A smaller lake draining into Hovvatn was not
limed and serves as a control.
9.2.2.3 United States
A paper by Pfeiffer (1982) reported results of a January 1981
questionnaire he circulated to Fisheries Chiefs or Directors in the
50 United States. Forty states responded to his acidic deposition •
questions. Nonrespondents included the states of Florida, Louisiana, •
Maryland, Michigan, Missouri, New Mexico, North Carolina, North
Dakota, Vermont and Virginia. Seven of 40 states replied that they •
are presently engaged in a liming program for ponded waters. The |
summary is as follows:
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West Virginia was the only state that indicated that they were liming •
streams. The figures provided were 16 km, representing approximately ™
12 ha. There were no questions on future considerations for liming
programs. •
Festa (pers. comm.) reported that the New York Department of
Environmental Conservation had treated 16 small ponds (0.5 to 3.0 ha) M
which were operated mainly as put-grow-and-take brook trout •
fisheries. The treated lakes had a simple food chain with only one
fish species stocked. Fish growth was good and the fish were
harvested by angling in the autumn. There was no attempt to •
establish a self-sustaining population. •
9.2.2.4 Ontario, Canada •
Limestone and slaked lime were added to Middle, Hannah, Lohi and
Nelson Lakes, four acidic lakes near Sudbury, Ontario between 1973 «
and 1976. Contamination by metals, especially Cu and Ni, prevented •
reestablishment of trout populations in the first three lakes which •
are situated within 13 km of Sudbury (Yan et al. 1979), even though
pH was increased from about 4.4 to >6.0. Nelson Lake (3.09 km2) •
was acid-stressed (pH -5.5-6.0) prior to additions of crushed |
limestone (51 metric tons) and slaked lime (68 metric tons) in the
fall of 1975 and the spring of 1976 (Yan et al. 1977). The decline f
of the lake's fisheries was indicated by the dominance of yellow •
perch and the disappearance of smallmouth bass. Lake trout
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populations were low but by the winter of 1979-1980 winter fishing
for lake trout was very successful (summarized in Yan and Dillon
1982).
9.2.3 Economic Aspects of Lake Liming
Estimating the total cost of liming aquatic systems is very difficult
but some unit cost figures are available for Sweden, New York State,
Norway and Nova Scotia. Generally there are three categories of cost
associated with liming programs: supply of chemicals, distribution
of chemicals, and monitoring of the systems before and after the
applications.
9.2.3.1 Costs in Sweden
Although the costs in Sweden cannot be expected to apply directly in
North America, they serve as a guide in estimating North American
costs. Results from their 5-year experimental program which dealt
with over 700 lakes (Bengtsson et al. 1980) have given good cost
estimates. The cost of limestone application, including materials
and distribution ranged from 500 to 1100 Skr ($115-253 Cdn.) per
metric ton using eight different spreading methods. The average cost
for the whole program was about $140 Cdn. per metric ton. Manual
applications had the lowest cost of about $115 Cdn. per metric ton
while aerial applications were the most expensive at about $250
Canadian per metric ton (National Fisheries Board and National
Environmental Protection Board 1981). The cost of scientific surveys
to document effects can range from a very small amount for some pH
and alkalinity measurements to several thousand dollars. In Sweden,
the average cost of research has been about $16,000 for each project.
However, each project may have more than one lake or river involved.
9.2.3.2 Costs in Norway
Limited cost information is available but a total experimental cost
of $80,000 for each of the five study lakes has been projected. In
addition there is support from universities with separate funding and
support from local residents.
9.2.3.3 Costs in New York State
New York State Department of Environmental Conservation has been
adding limestone to 16 small lakes (0.5 to 3 ha) and started a 40 ha
lake in 1979. They found the costs of limestone application to range
from $60 U.S. to $225 U.S. per hectare for a 3-year treatment ($20 to
$75 U.S./ha.yr). They conducted a very limited technical evaluation
of the lakes. The lakes were essentially devoid of fish to start
with and the objectives were to establish put-and-take brook trout
fisheries.
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9.2.3.4 Costs in Canada
9.2.4 Technical Evaluations Necessary in Liming Programs
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Watt (pers. comm.) has recorded actual cost figures for purchase, •
delivery and distribution of crushed bagged limestone to Sandy Lake |
in Nova Scotia. The loading rate was about one metric ton/ha. The
experiment was designed to protect salmon populations in downstream •
rivers. The lake was easily accessible, and crushed limestone was I
readily available. The total cost for liming the 70 ha lake was in
excess of $11,000, or about $160 Cdn. per hectare.
Although costs per lake will vary according to dose required, •
generally application rates for lakes appear to vary from about 380
(Yan and Dillon 1982) to 1000 kg/ha which Watt has used in Nova •
Scotia. The Ontario example given by Yan and Dillon (1982) has |
maintained the lake pH for at least five years. Generally, average
application rates of about 500 kg/ha are necessary to give multi-year .
pH stability. •
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The calculation of costs associated with neutralization programs must
be accompanied by the necessary technical evaluations. Any system, •
considered for liming must be studied in a variety of ways (e.g., •
depth, flushing time, water chemistry, and biota). Swedish treatment
and research costs of about $16,000 Cdn. per project are based on 304
projects which cover at least 700 individual lakes (Bengtsson et al. •
1980). Therefore, average monitoring costs appear to be about $8,000 •
per lake. A minimum sample program of only twice per year would
still cost at least $1,000 per lake including labour, analyses and •
data reporting. Meaningful evaluation of chemical and biological |
conditions would cost in the order of $10,000 per lake per year.
Control and management of the fisheries in treated systems would also •
add substantially to overall costs.
It is worth noting that the situation concerning fishing rights in I
Scandinavia and North America is quite different. In Sweden the •
rights to fish are privately owned, with the owners on some rivers
issuing fishing licenses and to some extent controlling fish harvest. •
This element of "self interest" allows for easier control of fishing •
activities which can affect the success of fish survival and repro-
duction. .
9.3 TERRESTRIAL LIMING
The addition of alkaline materials has been proposed as a means for •
ameliorating the effects of acidic deposition on terrestrial
ecosystems. While lime applications have an important place in the •
efficient management of agricultural soils and much research has been •
conducted to determine optimum dosages for different crops and soils,
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the scope for lime in temperate and boreal forestry is much more
limited. Moreover, few field trials have been concerned with entire
forested catchments. In this sub-section positive and negative
aspects of the liming technique will be discussed.
Numerous calcium-based alkaline materials are available for the
neutralization of acidified soil. However, for most situations,
crushed limestone (CaC03), flaked or hydrated lime (Ca(OH)2>, and
unslaked lime or quicklime (CaO) are the most readily available and
effective materials. A variety of substances have been proposed for
use as neutralization agents (Grahn and Hultberg 1975).
The data available up to now do not indicate obvious effects on
forest ecosystems caused by acid deposition. However, the potential
of the available techniques for remedial action warrant examination
in the event that subsequent data indicate forest degradation.
9.3.1 The Application of Lime to Agricultural Soils
Microorganisms and higher plants respond to their chemical environ-
ment, and soil kinetics are a key factor in determining agricultural
soil productivity. There are two major groups of factors which bring
about large changes in soil pH: (1) those which result in increased
adsorbed hydrogen and in turn release aluminum, and (2) those which
increase the content of adsorbed bases. Both organic and inorganic
acids are formed when organic matter is decomposed. The simplest and
perhaps the most widely found is carbonic acid (H2C03) which
results from the reaction of C02 and water. The solvent action of
H2C03 on the mineral constituents of the soil is exemplified by
its dissolution of limestone or calcium carbonate. Because carbonic
acid is relatively weak, it cannot account for the low pH values
found in many soils. Inorganic acids such as H2S04 and HN03
are suppliers of hydrogen ions in the soil. These acids, along with
the organic acids, contribute to the development of acid conditions.
Sulphuric and nitric acids are formed, not only by the organic decay
processes, but also from the microbial action on certain fertilizer
materials such as sulphur and ammonium sulphate. In the latter case
both nitric and sulphuric acids are formed.
Podzolization is an example of a process by which strong organic
acids are formed. The organic debris is attacked largely by fungi
which have among their important metabolic end products relatively
complex but strong organic acids. As these are leached into the
mineral portion of the soil, they not only supply hydrogen for
adsorption, but they also replace bases and encourage their solution
from the soil minerals. Leaching also encourages acidity.
Therefore, bases which have been replaced from the colloidal complex
or which have been dissolved by percolating acids are removed in the
drainage waters. This process encourages the development of acidity
in an indirect way by removing those metallic cations which might
compete with hydrogen and aluminum on the exchange complex.
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When lime is added to the soil, two changes occur:
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1) the calcium and magnesium compounds applied undergo solution •
under the influence of a variable partial pressure of carbon •
dioxide; and
2) an acid colloidal complex will adsorb considerable amounts |
of calcium and magnesium ions.
When lime, whether the oxide, hydroxide, or the carbonate, is applied I
to an acid soil, the movement, as solution occurs, is toward the ™
bicarbonate form. This is because the partial pressure of carbon
dioxide, usually several hundred times greater than that of •
atmospheric air, generally is intense enough to prevent the existence I
of the hydroxide or even the carbonate. The reactions, written only
for the purely calcium limes, are as follows: m
CaO + H20 ^a* Ca(OH)2
Ca(OH)2 + 2H2C03 ^=* Ca(HC03>2 + 2H20 I
H2C03 ^» Ca(HC03)2
The above equations represent only the solution of the lime in •
carbonated water. However, the soil situation is not as simple as
these reactions might lead one to assume. This is because the soil «
colloidal matter upsets the equilibrium tendencies by adsorbing the •
ions of calcium and magnesium. These ions may be taken from the soil
solution proper or directly from the solid phase if the contact is
sufficiently close (Buckman and Brady 1969). •
The changes of lime in the soil are many and complicated. If a soil
of pH 5.0 is limed to a more suitable pH value (e.g., pH 6.5) then a m
number of significant chemical changes occur. For example: (1) the •
concentration of H+ ions will decrease; (2) the concentration of
OH~ ions will increase; (3) the solubility of iron, aluminum and _
manganese will decline; (4) the availability of phosphates and •
molybdates will be augmented; (5) the exchangeable calcium and ™
magnesium will increase; (6) the percentage base saturation will
increase; and (7) the availability of potassium may be increased or
decreased, depending on soil conditions.
Overliming is an important phenomemon which must be considered. A ^
potential problem is the addition of lime until the pH of the soil is •
above that required for optimum plant growth. Under such conditions, *
many crops that ordinarily respond to lime are detrimentally
affected, especially during the first season following the lime I
application. With heavy soils, and when farmers can afford to apply •
only moderate amounts of lime, the danger is negligible. But on
sandy soils (low in organic matter and therefore lightly buffered) it
is easy to injure certain crops, even with a relatively moderate
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application of lime (Buckman and Brady 1969). Some of the
detrimental influences of excess lime are:
1. Deficiencies of available iron, manganese, copper, or zinc may
be induced.
2. Phosphate availability may decrease due to the formation of
complex and insoluble calcium phosphates.
3. The absorption of phosphorus by plants and especially its
metabolic use may be interfered with.
4. The uptake and utilization of boron may be hindered.
5. The drastic change in pH may, in itself, be detrimental.
9.3.2 Economics of Agricultural Liming
On judging the amounts of lime to apply, a number of factors should
be considered: (1) cost of liming material; (2) the soil surface pH,
texture and structure, and the amount of organic matter; (3) the
subsoil pH, texture and structure; (4) the crops to be grown; (5) the
length of the rotation; (6) the kind of lime used and its chemical
composition; (7) the fineness of the limestone; and (8) operational
experience.
9.3.3 Forest Liming
While much is known about agricultural liming practices (materials,
techniques, beneficial effects, and potential problems), much less is
known about liming forested ecosystems. For the boreal, north
temperate and temperate forests, such as are present in northeastern
North America, Scandinavia and northwestern Europe, liming has
considerable tradition for many centuries (Evelyn 1776).
Where forest liming is viewed more as a fertilizer or nutritional
measure, rather than as an aid to soil restoration, its promise is
far less re-assuring. This is because calcium deficiencies have
seldom been demonstrated and fertilizer trials embodying a calcium
treatment have rarely shown a positive response by tree growth.
Thus major reviews of fertilizer research for Canada (Rennie 1972),
the United States (Bengtsson 1977; Mustanoja and Leaf 1965), Sweden
(Holmen 1976), Great Britain (Everard 1974) and Germany (Baule and
Fricker 1970), show calcium trials to be extremely few compared with
those for nitrogen, potassium and phosphorus, with very few
indications of positive growth responses.
Forest liming has not been widely implemented because it has not yet
been shown statistically that acidic deposition has caused adverse
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effects in forest growth. Forest liming is further complicated
because the inaccessibility of forests makes application difficult
(Bache 1980). •
Sweden is one of the few countries where forest liming is practised.
The first attempts to lime forest lands in Sweden were made 67 years •
ago. The most recent lime applications were made at a rate of 5 - 10 I
metric tons/ha on 0.2 hectare plots in 12 areas. Various
combinations of fertilizer were also applied. It was pointed out
(Fraser et al. 1982) that after 25 years 50% of the lime was not I
leached from the soil. Research from 1971 to 1978 at Lisselbo •
(Fraser et al. 1982) where sulphuric acid and lime were applied to
plots, was described. Annual precipitation of 700 mm leaches between •
2.0% and 11.6% of the lime since application. Two preliminary |
conclusions from these studies are important: (1) liming has little
effect on the growth of forest trees and (2) lime persists in _
undisturbed forest soils, despite 700 mm of annual precipitation. •
Tveite and Abrahamsen (1980) report the results of field experiments
located in two different areas of southern Norway. The authors B
present results from the Norwegian field experiments with artificial |
acidic deposition and liming added to pine and spruce forests. All
experiments included treatment with 25 or 50 mm/month of artificial mm
acidic deposition with different pH, applied during the frost-free •
time of the year. After five years of treatment no negative growth
effects of the acid applications are apparent and there were no _
effects of liming found. •
No useful purpose would be served by documenting here a comprehensive
list of such trials, but a few typical published results exemplify •
the unattractiveness of the approach. For 45-year old jack pine |
(Pinus banksiana Lamb.) in the Boreal Forest of Ontario, calcium at
448 kg/ha gave no response except where nitrogen, phosphorus and •
potassium had also been applied (Morrison et al. 1977b). A further •
trial with 55-year old but poorer quality jack pine, also north of
Lake Superior, again only snowed a growth response to lime where
nitrogen, phosphorus and potassium has also been applied. Indeed, •
the suggestion was the lime by itself exercised a depressive effect •
upon growth by adversely affecting soil microbiological processes
(Morrison et al. 1977a). The complexity of lime effects is apparent •
from the work of Adams and his colleagues (1978) on the acid peaty £
gleys of Northern Ireland. There, lime did not increase the growth
of Sitka spruce (Picea sitchensis Carr.), but it did affect the soil _
microbiology and the viability of the mycorrhizal root association. I
As might be expected, the pH of the litter was raised from 4.0 to 6.0 •
- 6.5, a result that has been of serious concern to those aware of
the optimum soil conditions for the spread of rot fungi such as B
Formes annosus (de Azevedo and Moniz 1974). |
The possible effects associated with liming forested ecosystems are •
still unknown but experiments of watershed liming may provide some •
insight. Bengtsson et al. (1980) report on experiments of watershed
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liming conducted in Sweden. Agricultural lime (i.e., powdered
CaC03) is generally transported to the watershed in large trucks
and applied as a slurry with a sprayer truck. The CaCo3 dose
required to achieve adequate neutralization of watershed systems is
generally two orders of magnitude greater than that of direct water
addition which is due to the many base consuming processes that occur
within the forest soil systems (Bengtsson et al. 1980). There have
been application rates reported in the range of 5,000-7,000 kg
CaC03/ha.yr. Hultberg and Andersson (1982) reported that some
damage to the terrestrial environment may be associated with liming.
Sphagnum moss was severely damaged with CaC03 addition. Damage to
lichens, mosses and spruce needles was also observed.
Smelters at Sudbury, Ontario, represent the greatest single source of
sulphur dioxide emission in the world. Over the past several years,
a terrestrial liming/reclamation program has been operated (Fraser
et al. 1982). The affected lands have a pH of approximately 4.0 and
concentrations of copper and nickel were measured up to 10 ppm. To
reclaim this land, crushed agricultural limestone is applied at a
rate of 12.4 short tons/ha, then fertilized with a nitrogen-
phosphorus-potash mixture (6-24-24, respectively), and seeded with a
variety of blended grasses. The limestone application is labor
intensive with 400 students adding the limestone by hand. Over 1000
hectares have been reclaimed to date. According to the authors, this
terrestrial liming project has been extremely successful in raising
the pH of the soil and complexing the heavy metals (although there is
still some minor nickel toxicity). The authors report that grass is
now able to grow on barren areas and the resultant shading and
lowering of ground temperature has enabled some natural vegetation
(e.g., quaking aspen seedlings) to become reestablished. The newly
established vegetation is monitored and analyzed, as is the recovery
of populations of insects, birds, and small mammals.
Limited information is available on the changes to terrestrial flora
and fauna after the addition of lime to forested soils. Tamm (1976)
stated that when lime was added to forest soils in small-scale
experiments, tree growth rates typically was not enhanced, because of
the tendency of lime to immobilize the nitrogen in organic matter and
thereby reduce its availability to trees. Fraser et al. (1982)
report that lime had little effect on the growth of forest trees,
based on preliminary conclusions from forest liming studies conducted
in Sweden from 1971 to 1978. Abrahamsen et al. (1980) reported that
soil animal populations nearly always fail to increase when soil
acidity was reduced by liming. Hultberg and Andersson (1982) report
that the liming of watersheds in addition to lakes and streams would
release additional phosphorous to the waterbodies and enable these
aquatic systems to increase their primary productivity.
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9.3.4 Terrestrial Liming Summary
9.4 DRINKING WATER SUPPLY
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In conclusion, although in principle the liming of land might I
neutralize an acidifying pollutant (Ulrich 1972) it has the following I
serious limitations. First, it would not prevent direct injury to
plant tissues, even where in agricultural situations it is already •
being used as a soil amendment. Secondly, in the typical forest •
situation it would be very difficult to apply. Thirdly, the effects
of lime in the boreal and north temperature forest are complex and
often far from beneficial. •
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Low pH conditions in municipal water supplies can cause corrosion of
the plumbing materials. The estimated costs for controlling «
corrosion were based on adding lime to water to stabilize it. This •
should control corrosion of lead pipes as well as corrosion of cast
iron water mains, and is the corrosion control technique most likely
to be used by water utilities. I
9.5 COSTS OF CORROSION CONTROL M
Costs for corrosion control by lime stabilization were estimated by
Hudson and Gilcreas (1976) to total $0.30 U.S. per capita for
operation and amortization costs in 1976. Average per capita costs •
for lime stabilization were estimated by Davis et al. (1979) to *
range from $0.18 to $0.57 U.S., depending on the extent of chemical
treatment provided to stabilize the water.
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Corrosion control by calcium carbonate stabilization and deposition
of a protective calcium carbonate film has been suggested by EPA as •
an effective approach to nonselectively provide protection to a •
number of materials, including asbestos cement, lead, iron,
galvanized steel, copper, and alloys that may be used in water
distribution systems or plumbing. Annual per capita costs for •
corrosion control by addition of lime were estimated by USEPA (1979). •
Costs are a function of plant size, as shown in Table 9-1.
The methods used to calculate corrosion costs in the USEPA Statement J|
of Basis and Purpose were developed by Gumerman et al. (1979). An
example of calculation of cost for corrosion control by addition of _
lime at 30 mg/L for pH control in a 5 million gallons per day (MGD) •
plant is given in Table 9-2. The costs are shown for operation at •
70% of capacity (3.5 MGD). The principle items of expense are
capital amortization, labor, and chemical used. Chemical consumption I
is the cost category most sensitive to water quality changes. |
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TABLE 9-1. COST OF CORROSION CONTROL BY LIME ADDITION AS A FUNCTION
OF PLANT SIZE
Plant Size
Gallons Per Day
2,500
50,000
5,000,000
100,000,000
(t/ 1,000 Gallons
Treated - 70% Capacity
56.9
16.4
2.7
1.1
Annual Per Capita
Costa - $
20.50
6.00
1.00
0.40
Cost stated in $ U.S. 1981, December.
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TABLE 9-2. EXAMPLE OF COST CALCULATION FOR FEEDING LIME AT A
5 MILLION GALLONS PER DAY PLANT
$/Yeara
a Cost per 1,000 gallons treated = 2.68
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9.6 REFERENCES
Abrahamsen, G.; Hovland, J.; and Hagvar, S. 1980. Effects of
artificial rain and liming on soil organisms and the
decomposition of organic matter. In Effects of acid
precipitation on terrestrial ecosystems, eds. T.C. Hutchinson
and M. Havas, pp. 341-362. New York: Plenum Press.
Adams, S.N.; Cooper, J.E.; Dickson, E.L.; and Seaby, D.A. 1978.
Some effects of lime and fertilizer on a Sitka spruce
plantation. Forestry 51(1):57-65.
Bache, B.W. 1980. The acidification of soils. In Effects of acid
precipitation on terrestrial ecosystems, eds. T.C. Hutchinson
and M. Havas, pp. 183-202. New York: Plenum Press.
Baule, H., and Fricker, C. 1970. The fertilizer treatment of forest
trees. Munich: BLV-Verlag.
Bengtsson, B. Personal communication. Swedish Water and Air
Pollution Research Institute (IVL), Goteborg, Sweden.
Bengtsson B.; Dickson, W.; and Nyberg, P. 1980. Liming acid lakes
in Sweden. Ambio 9(l):34-36.
Bengtsson, G.W. 1977. Fertilizers in use and under evaluation in
silviculture; a status report. International Union of Forestry
Research Organizations, XVI World Congress, Oslo. 32 pp.
Buckman, H.D., and Brady, N.C. 1969. Nature and properties of
soils. New York: McMillan Co.
Davis, M.J., et al. 1979. Occurrence, economic implication and
health effects associated with agressive waters in public water
systems.Final Report, MRI Project No. 4552-L, Prepared for A/C
Type Producers Association.
de Azevedo, N.F., and Moniz, M. 1974. Influence of temperature, pH
and nutrients on the growth of Fomes annosus isolates. In
Fourth Int. Conf. on Fomes Annosus, pp. 163-168. IUFRO, Section
24, Athens, Georgia, USDA Forest Service, Washington, DC.
Evelyn, J. 1776. Silya, a discourse of forest trees, new edition,
with Notes by 7^. Hunter. York.
Everard, J.E. 1974. Fertilizers in the establishment of conifers in
Wales and southern England.Forestry Commission Booklet,
No. 41, HMSO, London. 49 pp.
Festa, P.J. Personal communication. Department of Environmental
Conservation, Albany, NY.
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Fraser, J.; Hinckley, D.; Burt, R.; Severn, R.R.; and Wisniewski, J.
1982. A feasibility study to utilize liming as a technique to
mitigate surface water acidification. EA-2362 Final Report,
Electric Power Research Institute, Palo Alto, CA.
Hudson, H.E., Jr., and Gilcreas, F.W. 1976. Health and economic
aspects of water hard
Assoc. 68(4):201-204.
1981. (in press)
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Grahn, 0., and Hultberg, H. 1975. The neutralizing capacity of 12 «
different lime products used for pH-adjustment of acid water. •
Vatten 2:120-21. *
Gumerman, R.C.; Gulp, R.L.; and Hansen, S.P. 1979. Estimating water •
treatment costs. Volume 2. Cost curves applicable to 1 to 200 •
mgd treatment plants. Final Report EPA-600/2-79-162B, U.S.
Environmental Protection Agency, Santa Ana, CA. •
Holmen, H. 1976. Forest fertilization in Sweden 1975. Kungl.
Skogs-och Lantbruksakadamiens Tidskrift 1976(3) Medd. No. 7. _
9 PP.
aspects of water hardness and corrosiveness. J. Am. Water Works^ •
Hultberg, H., and Andersson, I.E. 1982. Liming of acidified lakes: •
induced long-term changes. Water, Air, Soil Pollut. 17. (in •
press)
Morrison, I.K.; Swan, H.S.D.; Foster, N.W.; and Winston, D.A. 1977a. I
Ten-year growth in two fertilization experiments in a semi-
mature jack pine stand in northwestern Ontario. For. Chron.
53:142-146. •
Morrison, I.K.; Winston, D.A.; and Foster, N.W. 1977b. Effect of
calcium and magnesium, with and without NPK on growth of semi- •
mature jack pine forest, Chapleau, Ontario: fifth-year results. •
Can. For. Serv. Report O-X-259, Sault Ste Marie, Ont. 11 pp.
Mustanoja, K.J., and Leaf, A.L. 1965. Forest fertilization •
research, 1957-1964. Bot. Rev. 31(2):151-246. •
National Fisheries Board and National Environmental Protection Board.
1981. Liming of lakes and rivers, 1977-1981, in Sweden, eds.
B. Bengtsson, and L. Henriksson, Goteborg, Sweden.
Pfeiffer, M.H. 1982. Recent impacts of acidification on fisheries •
resources in the U.S. In Proc. Int. Symp. Acidic
Precipitation and Fishery Impacts in Northeastern North America.
American Fisheries Society, Cornell University, Ithaca, NY., •
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